Nitrogen in boreal forest ecosystems

Transcription

1 APPENDIX B Nitrogen in boreal forest ecosystems The influence of climate change, nitrogen fertilization, wholetree harvest and stump harvest on the nitrogen cycle A report from Belyazid Consulting & Communication

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3 A report from Belyazid Consulting and Communication AB This report is an appendix to the report Effects of climate change, nitrogen fertilization, wholetree harvesting and stump harvesting on boreal forest ecosystems A review of current knowledge and an evaluation of how these factors may influence the possibilities to reach the Swedish environmental objectives. January Authors Ulrika Jönsson Belyazid & Salim Belyazid, Belyazid Consulting and Communication AB, Fersens väg 9, Malmö, Sweden Cecilia Akselsson, Physical Geography and Ecosystem Sciences, Lund University, Sölvegatan 12, Lund, Sweden Contact Ulrika Jönsson Belyazid, ulrika@belyazid.com, Tel: Acknowledgements This work was funded by the research program Climate Change and Environmental Objectives (CLEO) hosted by IVL Swedish Environmental Research Institute and the strategic research area Biodiversity and Ecosystem Services in a Changing Climate (BECC) hosted by Lund University. The photos in the report are from dreamstime.com, freedigitalphotos.net and shutterstock.com. The photo on page 18 was kindly provided by Rasmus Kjoller.

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5 SUMMARY This report consists of a literature review about the effects of climate change and forest management on nitrogen (N) cycling in boreal forest ecosystems, with specific emphasis on Swedish conditions and the terrestrial part of the N cycle. With regard to climate change, the influence of temperature, moisture and increased levels of carbon dioxide (CO 2 ) on N cycling have been evaluated. With regard to forest management practices, the effects of N fertilization, wholetree harvesting (WTH) and stump harvesting (SH) have been evaluated. Results from empirical as well as modelling studies are included in the review. N input N is mainly added to terrestrial ecosystems through biological N fixation and atmospheric deposition. In production forests, it may also be added in the form of N fertilizer. Previously, boreal forests have been considered to have a limited capacity for N fixation. However, recent studies have shown substantial levels of N fixation by cyanobacteria colonizing pleurocarpus feather mosses in northern boreal forests, and this process is now considered to be the main source of N in these ecosystems. Since microbial activity generally increases with temperature, and the efficiency of the Nfixing enzyme nitrogenase reaches its maximum near 25 C, an increasing mean annual temperature is likely to have a positive effect on Nfixation rates something that also has been demonstrated in a number of studies in northern Sweden. However, the high light sensitivity of feather mosses could mean that climateinduced increases in plant productivity could eventually limit biological N fixation because of greater shading. N fixation is also generally stimulated by increased moisture (nitrogenase is very sensitive to oxygen) and by elevated CO 2 (although the number of studies in boreal forests are limited). It is also well established that Nfixation rates are directly influenced by soluble N concentrations. N additions to a latesuccession forest in the north of Sweden was found to have dramatic and lasting influence on Nfixation rates and fertilization at a rate of 4.5 kg N ha 1 yr 1 basically eliminated any N fixation throughout the study period. Many other studies show similar results. The effects of WTH and SH on N fixation are basically unknown. With regard to the deposition of N, human activities have led to major increases in the global emissions of N to the atmosphere and, subsequently, to increases in the deposition of biologically available N to the terrestrial biosphere. In Sweden, there is a clear gradient, with wet deposition of inorganic N decreasing substantially from the southwest ( 11 kg ha 1 yr 1 ) to the north ( 2 kg ha 1 yr 1 ). In addition, there may be substantial amounts of N reaching ecosystems by dry deposition. Fertilization with N has been used in Sweden as a measure to increase tree growth since the 1960s. Around 2 million ha of forest has been subjected to N fertilization at some occasion. After a peak in the 1970s ( ha yr 1 fertilized), the area being fertilized decreased to reach its lowest points between 1995 and 2005 ( ha yr 1 ) and has then increased again ( ha yr 1 in 2009). N in plants The plant Nuptake rate depends on the concentration of N in the environment as well as on the demand of the plant, the latter being at least partly determined by the plant growth rate. It is wellknown that plants have several different mechanisms to enhance the availability of nutrients in the rhizosphere. With regard to N, mycorrhizal associations are one of the most important. That N uptake increases with temperature, at least within the temperature ranges that are observed in boreal forest ecosystems, have been demonstrated in both laboratory and field studies. The impact of drought and excess water availability is less certain. However, it seems likely to assume that drought results in an impaired N uptake, since extended drought periods leads to reduced fine root biomass and may also influence the nutrient availability and transport in the soil. Increased levels of CO 2 generally result in an increased allocation of C to belowground parts of trees, with subsequent increases in the growth of roots and mycorrhiza, which should allow the trees to exploit a larger soil volume and thus might be expected to promote N uptake. Indeed, more often than not, an increase in N uptake seems to occur at elevated CO 2, provided there is an adequate supply of N in the soil. The increase is most probably a function of increased root growth, since most studies to date have shown little effect, or no effect at all, of elevated CO 2 on soil Nmineralization rates. There are basically no scientifically published studies on how WTH and SH may affect N uptake and acquisition of trees in the subsequent generation. The two studies that do exist are greenhouse studies. They show that N in brash and in decomposing roots can make small, but discernible, contributions to new tree growth after one growing season. Elevated temperatures are generally believed to increase the concentrations of N in plant tissue, because mineralization and diffusion of N are stimulated at high temperatures. However, the response probably depend on the nature of the temperature increase. While a small and gradual increase in mean annual temperature may give the above mentioned results, a period of very high seasonal maximum is likely to have the opposite effect, not least because of its interaction with drought. Drought generally causes an impairment of the nutrient concentrations and contents of plants. Modelling results suggest that this may be a consequence of lower decomposition rates and, consequently, lower availability of N. It may also be a consequence of reduced C allocation to belowground, resulting in less C and energy available for uptake and assimilation. The initial stimulation of photosynthesis under elevated CO 2 is often accompanied by a decrease in plant N concentration, in particular foliar N and especially at low N supply. The decrease has been suggested to be a result of either decreased

6 N uptake per unit of biomass produced or preferential allocation of acquired N to root tissue. Fertilization experiments with N have generally shown increased leaf N concentrations, although the increase may sometimes be rather shortlived. Studies of WTH, on the other hand, usually show no effect on N concentrations of trees (in particular the longterm ones). The effects of SH on N concentrations of trees in the subsequent generation have rarely been studied, but the two studies available showed lower concentrations of N in SH stands as compared with conventionally harvested (CH) stands (despite N having been added before planting in one of the two studies). The quantity of N transferred from the trees to the soil by litterfall is primarily a function of tree biomass. Generally, increases in temperature, moisture, CO 2 and N are all regarded to result in increased production and thus litterfall. WTH and SH, on the other hand, implies a removal of potential litter, and thus N, from the ecosystems. N in soils Rates of decomposition, mineralization and nitrification generally increase with temperature, at least under favourable moisture conditions and in ecosystems where temperature is a limiting factor. The effect of drought on mineralization rates in forest ecosystems is still poorly understood and results are variable. Ultimately, however, reduced soil water availability limits the microbial activity in soils, and it should thus also limit mineralization. With regard to nitrification, it is generally inhibited at very low soil moisture levels, increases with soil moisture to an optimum and then decline as the soil becomes water saturated. Freezingthawing and dryingrewetting cycles have both been demonstrated to occasionally result in increased N mineralization and nitrification rates. However, the increases, if observed, are usually rather small in relation to the expected annual rates. Most studies to date have indicated little or no effect of elevated CO 2 on soil Nmineralization rates. The results with regard to nitrification are inconclusive, with studies reporting increases, decreases and no change in response to elevated CO 2. Gross N mineralization usually shows a strong negative correlation with the substrate C/N ratio in longterm studies. The increase in mineralization in response to increased availability of N is probably a result of increased plant productivity, but also of an increase in the activity and biomass of the soil microbial community. Possibly, it may also be an effect of a change in the composition of the soil microbial community at high levels of N. N availability is also a major controlling factor of nitrification, and many studies have reported increased nitrification rates in response to N addition, especially in ecosystems experiencing elevated N inputs over prolonged periods of time. Since WTH and SH result in reduced substrate availability (and probably also influence the abiotic environment), it seems logical to assume that they should influence the processes of decomposition, mineralization and nitrification and, eventually, result in diminishing stores of N in forest soils. With regard to SH, the few studies available do indeed report significantly reduced soil N pools and mineralizable N after SH as compared with CH. This is also a result commonly found for WTH in nutrient budget calculations and when using dynamic modelling. In reality, however, significant differences between WTH and CH are rarely discernible in the field, at least in the Scandinavian countries. Considering SH, keeping stumps were recently suggested to constitute an important mechanism of nutrient retention, potentially diminishing nutrient leaching and maintaining site fertility and productivity. N losses N is lost from ecosystems mainly through denitrification, leaching and harvesting. However, N mineralisation, N fixation and nitrification also contribute to the losses of volatile N compounds. Despite the theoretical temperature dependence of denitrification, the temperature dependence in the field is usually not pronounced. This is probably due to the indirect effects of temperature (i.e. low temperature enhances soil oxygen consumption, inducing anoxic conditions which promote denitrification) or the fact that other environmental controls exert a stronger influence on denitrification than does temperature. Denitrification is enhanced by anoxic conditions and high soil water content thus usually result in enhanced denitrification rates. Denitrification is also strongly dependent on C availability and there exists a general correlation between total soil organic matter content and denitrification potential. The increase in C allocation belowground as a consequence of elevated CO 2 is thus expected to result in a stimulation of denitrification. However, conclusive field results are missing. Most studies to date have demonstrated that N additions generally lead to increased fluxes of gaseous N from the soil, although the emissions are not necessarily high or correlated to the amount of N added. To our knowledge, there are no studies on how WTH and SH influence denitrification. There is still relatively little information available with regard to how various environmental factors affect the leaching of N. Primarily, leaching is limited by the availability of soluble N. Secondly, hydrology imposes a strong control on the transport of N in soil, with increasing precipitation being associated with more rapid leaching. Although increased temperature generally results in enhanced mineralization rates, this does not necessarily lead to additional N losses, as N uptake by roots or immobilization by microbes might occur. However, the potential risk of leaching increases. Studies of N leaching from forests subjected to elevated CO 2 are rare, and the results variable. Several fertilization studies have indicated that leaching may be induced when the availability of N increases. However, leaching only occurs if the N input exceeds the capacity of the ecosystem to retain N, something that is generally the case only at sites with very large N inputs. Recent studies in several European countries have indicated that N leaching is one of the last things to occur in the succession of events related to increased N availability in ecosystems. Several studies have suggested that removal of logging residues may reduce the accumulation of N in forest ecosystems and thus the potential leaching of N. Although there are some studies supporting this statement, the fieldbased evidence is relatively scarce and there are studies (both field and modelling ones) which have reported the opposite result, i.e. a decreased leaching after CH as

7 compared with WTH. The effects of SH on leaching is unclear, since the number of studies is very low. The amount of N lost through harvesting depends not only on the amount of harvested biomass, but also on what fraction of the tree that is harvested. Nutrient budget calculations have shown that WTH may result in up to two times higher N loss than CH. Even if the N losses are smaller when stumps are removed (since N concentrations are lower than in slash), they are not negligible since the biomass of stumps is substantial. Conclusions Although recent advances have resulted in a more comprehensive understanding of how climate change and forest management practices affect the processes controlling the N dynamics of boreal forest ecosystems, much information is still lacking. In particular, the longterm impact on soil N and N losses from ecosystems are unclear. Improving our knowledge about soil N, endeavouring in obtaining longterm data sets, and clarifying the couplings between the N and the C cycles are crucial if reliable predictions about the future cycling of N in boreal forest ecosystems are to be made.

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9 TABLE OF CONTENTS SUMMARY BRIEF NITROGEN CYCLING AN OVERVIEW NITROGEN INPUTS N fixation Effects of temperature Effects of moisture Effects of CO Effects of N Effects of WTH Effects of SH Atmospheric deposition of N Throughfall N fertilization NITROGEN IN PLANTS N acquisition and uptake Effects of temperature Effects of moisture Effects of CO Effects of N Effects of WTH Effects of SH N content of trees, allocation, retranslocation and storage Effects of temperature Effects of moisture Effects of CO Effects of N Effects of WTH Effects of SH Plant growth and aboveground production Root growth and mycorrhiza production Litterfall Effects of temperature Effects of moisture Effects of CO Effects of N Effects of WTH Effects of SH NITROGEN IN SOILS Decomposition, mineralization and immobilization Effects of temperature Effects of moisture Effects of CO Effects of N Effects of WTH Effects of SH Nitrification Effects of temperature Effects of moisture Effects of CO Effects of N Effects of WTH Effects of SH NITROGEN LOSSES Denitrification and volatilization Effects of temperature Effects of moisture Effects of CO Effects of N 33

10 6.1.5 Effects of WTH Effects of SH Leaching Effects of temperature Effects of moisture Effects of CO Effects of N Effects of WTH Effects of SH Tree harvest CONCLUSIONS REFERENCES

11 1. BRIEF This appendix consists of a literature review about the effects of climate change and forest management on nitrogen cycling in boreal forest ecosystems, with specific emphasis on Swedish conditions. The literature review is an appendix to the report Effects of climate change, nitrogen fertilization, wholetree harvesting and stump harvesting on boreal forest ecosystems A review of current knowledge and an evaluation of how these factors may influence the possibilities to reach the Swedish environmental objectives published by Belyazid Consulting & Communication AB in January The aims of this review are: To identify what processes are controlling the N dynamics in boreal and temperate forest ecosystems. To investigate how these processes are influenced by climate change and some forest management practices that are likely to become more common in Sweden in the future, namely nitrogen (N) fertilization, wholetree harvesting (WTH) and stump harvesting (SH). With regard to climate change, the influence of temperature and moisture as well as increased levels of carbon dioxide (CO2) on various processes in the N cycle are evaluated. The focus of the literature review is on the terrestrial part of the N cycle. The processes included are: N fixation, N acquisition and uptake of plants, plant growth, litterfall, decomposition, mineralization, immobilization, nitrification, denitrification and leaching. We are aware of the difficulty in drawing coherence from a large number of studies, which to varying degrees, are site, age and speciesspecific. However, we have tried to include as much information as possible within the limited time available. Foremost, we have collected information from scientifically published articles concerning boreal forest ecosystems in Scandinavia. When the information available was scarce, we extended the literature search to other types of forest ecosystems and other regions of the world. Both empirical and modelled data have been included in the review. For SH, very little scientifically published material is available. Consequently, the conclusions with regard to the effects of SH on forest ecosystems are uncertain. 9

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13 2. NITROGEN CYCLING AN OVERVIEW The overview is mainly based on the description provided by Brady & Weil (1999). As it moves through the N cycle, an atom of N may appear in many different chemical forms, each with its own properties, behaviours, and consequences for the ecosystem. The principal pools and forms of N, and the processes by which they interact in the cycle, are briefly described below. For more detailed information about the major pathways and processes governing N in boreal forest ecosystems, see sections three to six. In the atmosphere, N exists mainly as dinitrogen gas (N 2 ), with traces of nitric oxide (NO), nitrogen dioxide (NO 2 ), nitrous oxide (N 2 O), and ammonia (NH 3 ). N in the form of N 2 is and has always been abundant, making up about 78% of the world s atmosphere. N 2 is, however, unavailable to plants. It has to be fixed, i.e. pulled from the air and bonded to hydrogen or oxygen to form inorganic compounds, mainly ammonium (NH 4 ) and nitrate (NO 3 ), before the plants can use it. This conversion is performed naturally either by lightning or by certain microorganisms (including some species of bacteria, a number of actinomycetes and cyanobacteria). Globally, enormous amounts of N are fixed each year terrestrial ecosystems alone fix an estimated 139 million Mg N per year, while lightning fixes around 8 million Mg N per year (Brady & Weil, 1999). Besides this natural fixation, human activities during the past century have resulted in substantial increases in the global N fixation, effectively doubling the annual transfer of N from the atmospheric pool to biologically available forms (Vitousek et al., 1997). The major sources of this enhanced supply include industrial processes that produce N fertilizers, the combustion of fossil fuels, and the cultivation of soybeans, peas, and other crops that host symbiotic Nfixing bacteria. As a consequence, substantial amounts of N nowadays also enter the ecosystem as atmospheric deposition (NH 3, NH 4 or NO 3 ) or, in production forests, in the form of N fertilizer. Although biological fixation is the most important process for the introduction of combined N into natural environments on a global scale (Son, 2001), deposition and fertilization may be of greater significance locally. Once N has entered a terrestrial ecosystem, it undergoes a complex series of alterations. Ammonia gas may be absorbed directly by plant foliage (Hutchinson et al., 1972), whereas NO 3 and NH 4 ions enter the ecosystems as dry fallout or in precipitation, and is taken up either by leaves or roots of trees. The ions are then converted to organic N, in a process known as assimilation, before proceeding along either grazing or detritus trophic chains (decomposition) to end up as soil N. The soil is the major sink for N in most ecosystems, and Kauppi et al. (1995), Tietema (1998) and Nadelhoffer et al. (1999) showed that between 70 and 90% of the N added to forest ecosystems in 15 N experiments ended up in the soil compartment. Initially, the main sink is the forest floor, but later, most of the 15 N is recovered from the lower mineral soil (Emmett et al., 1995). The great bulk (95 to 99%) of the soil N is in organic compounds (Brady & Weil, 1999). Organic N may be stored in soils for many centuries, or even millennia (Kimmins, 1997). Sooner or later, however, much of it is converted to NH 4 by the process of mineralization. The NH 4 ions are subject to five possible fates in the N cycle (Brady & Weil, 1999). They may be: 1) removed by plant uptake, 2) immobilized (i.e. converted into organic form again) by microorganisms, 3) fixed in the interlayers of certain clays, 4) transformed into NH 3 and lost to the atmosphere by volatilization or 5) oxidized to nitrite (NO 2 ) and subsequently to NO 3 by the microbial process of nitrification. Nitrogen in the NO 3 form is highly mobile in the soil and in the environment and NO 3 released by nitrification is thus quickly: 1) taken up by plants, 2) immobilized by soil microbes or 3) lost by leaching in drainage water. It may also be 4) volatilized to the atmosphere by denitrification, a process in which NO 3 is reduced to NO, N 2 O, N 2 or NH 3. The rates of the various steps included in the belowground part of the N cycle depend mainly on temperature and soil conditions. At each step, plants or microbes can take up soluble N, or N can be leached from the system, reducing the substrate available for the next N transformation. Therefore, the supply rate of different forms of available N to plants and microbes follow the sequence: dissolved organic N NH 4 NO 3 (Lambers et al., 1998). However, although the N supply rate always follows the same sequence in soils, the quantities and relative concentrations of these soluble forms of N varies substantially. That the recycling of N from plants to soil and back to plants is a key component regulating the N availability in soil and hence the rate of forest growth is evident from the fact that approximately 93% of the N that is available for plant growth in temperate forests originates from this type of recycling, while only 7% comes from the atmosphere (Lambers et al., 1998). In arctic tundras, the number is even higher reaching 96% (Lambers et al., 1998). 11

14 3. NITROGEN INPUTS N is mainly added to terrestrial ecosystems through biological N fixation and atmospheric deposition. However, in production forests, N may also be added to the ecosystem in the form of N fertilizer. Among these three input forms, biological fixation is the most important process for the introduction of combined N into natural environments on a global scale, although deposition and fertilization may be of greater significance locally (Son, 2001). 3.1 N fixation Previously, boreal forests have been considered to have an extremely limited capacity for N fixation, with fixation rates estimated at only 0.5 kg N ha 1 yr 1 (DeLuca et al., 2002). However, recent studies have shown substantial levels of N fixation by cyanobacteria colonizing pleurocarpus feather mosses (DeLuca et al., 2002), and N fixation is now considered as the main source of ecosystem N in northern boreal forest ecosystems (DeLuca et al., 2008). The fixation rates in feather moss carpets may, however, vary greatly, both spatially and in terms of time (DeLuca et al., 2002) as well as with stage of secondary succession (Zackrisson et al., 2004). For earlysuccession sites (2580 years since fire), Zachrisson et al. (2004) reported N fixation rates below 0.5 kg ha 1 yr 1, while midsuccesion sites ( years since fire) had fixation rates between 0.4 and 1.6 kg ha 1 yr 1 and latesuccession sites (>200 years since fire) had fixation rates between 1.0 and 2.0 kg ha 1 yr 1. Recently, N fixation rates of up to 4 kg N ha 1 yr 1 in boreal forests were reported by Gundale et al. (2010; 2011). Despite the importance of biological N fixation in boreal ecosystems, very little is known about how this process may respond to climate change (Gundale et al., 2012). In addition to symbiotic fixation of atmospheric N, asymbiotic fixation may also contribute to the input of N in forest ecosystems (Son, 2001). Asymbiotic fixation is known to occur in decaying wood, in the forest floor, in the rhizosphere, by bacteria associated with mycorrhizal fungi and even in the stemwood of living trees (Kimmins, 1997). Although the rates are quite low (generally less than 3 kg ha 1 yr 1 and mostly less than 1 kg ha 1 yr 1 ), this addition has been suggested to play an important role in maintaining N ecosystem budgets over long time periods (Kimmins, 1997; Son, 2001) Effects of temperature In general, microbial activity increases with temperature (Brady & Weil, 1999). According to Gundale et al. (2012), it is also known that the efficiency of the Nfixing enzyme nitrogenase reaches its maximum near 25 C. An increasing mean annual temperature is thus likely to have a direct positive influence on Nfixation rates. However, it is also possible that warmer temperatures may enhance evaporative water loss, which have been shown to affect feather moss Nfixation rates negatively (Gundale et al., 2009; Jackson et al., 2011). A recent study by Gundale et al. (2012), investigating the interactive effects of temperature and light on biological N fixation in boreal forests, showed that warming enhanced Nfixation rates of the two feather mosses investigated Pleurozium schreberi and Hylocomium splendens, with the former responding more strongly to the temperature increase than the latter. However, for both species, the highest warming treatment (30.3 C) eventually caused N fixation to decline. Light strongly interacted with warming treatments, having a positive effect at low or intermediate temperatures but damaging effects at high temperatures. That temperature has a positive effect on Nfixation rates and that the temperature sensitivity may differ between species was also shown in a field experiment by Markham (2009), examining Nfixation rates in wet lowland and dry upland boreal forests in central Canada. Based on their results, Gundale et al. (2012) concluded that the maximum predicted increases in mean annual temperatures (i.e. 5 C; IPPC, 2007) during the next century is likely to directly increase Nfixation rates by the two feather moss species investigated, meaning that increased biological N inputs may satisfy the higher N requirements of a more productive vascular plant community in warmer climates. However, Gundale et al. (2012) cautions that N fixation by the two mosses are also sensitive to light availability, which could mean that climateinduced increases in plant productivity could limit biological N fixation because of greater shading. Furthermore, exposure to extreme temperature events substantially damaged the Nfixation capacity of the feathermoss species, suggesting that increases in the frequency and intensity of extreme temperature events in the future could reduce biological N inputs in some boreal environments Effects of moisture The Nfixing enzyme, nitrogenase, is very sensitive to oxygen. Generally, N fixation is thus stimulated by reduced oxygen levels and low redox potentials (Paul & Clark, 1996; Son, 2001). Soil moisture levels are thus linked to N fixation through its influence on oxygen supply, but also through its influence on transportation of mineral nutrients, many of which are essential for N fixation to occur (P, Mo, Fe, Co, Mg, P, W and Ca) (Son, 2001; Vitousek et al., 2002). Recent studies on the influence of variations in moisture and rainfall intensity and frequency in boreal forests in northern Sweden have verified the strong dependency of N fixation on soil water availability. Gundale et al. (2009) found that prolonged drought (>45 days) resulted in a diminished Nfixation capacity by the feathermosscyanobacteria association investigated, whereas persistent moisture resulted in an increase in the Nfixation capacity. The sensitivity to moisture was larger in feathermosses collected from old forests than from young forests. Gundale et al. (2009) thus concluded that moisture availability may be an important regulator of Nfixation rates in some boreal forest environments. Their results were supported by those of Jackson et al. (2011), who showed that N fixation of boreal feather mosses under greenhouse conditions was strongly influenced by both rainfall amount and 12

15 frequency, as well as their interaction (an increase in the frequency of rain had greater effect when the amounts were higher). Jackson et al. (2011) also found that feather mosses had a buffering effect on the decomposer subsystem under moisture limitation, promoting vascular litter decomposition rates, concentrations of litter nutrients and active soil microbial biomass and reducing N release into soil solution under low soil moisture conditions. Several studies performed in nonboreal systems have also demonstrated correlations between moisture and N fixation in bryophytes or lichens (Turetsky, 2003; Zielke et al., 2005) Effects of CO2 According to a review on longterm responses of forest ecosystems to elevated CO2 (Johnson, 2006), elevated CO2 has been found to stimulate symbiotic N fixation in trees in several studies (see for example Norby, 1987; Arnone & Gordon, 1990). Arnone & Gordon (1990) found evidence for the presence of a positive feedback loop between N fixation and photosynthesis in nodulated Alnus rubra plants growing at elevated CO2. The plants had greater wholeplant photosynthesis, nitrogenase activity, leaf area, N content, and nodule and plant dry mass relative to nodulated plants grown at ambient CO2 and nonnodulated plants grown at both CO2 levels. The relative amount of dry mass allocated belowground decreased for all seedlings over time and the amount allocated aboveground increased. The proportion of dry mass allocated belowground was consistently greater in nonnodulated plants. However, a stimulation of symbiotic N fixation does not occur in every case (Arnone, 1999). tion treatment (3 kg N ha1 yr1) resulting in a nearly 50% reduction in N fixation per unit mass relative to control. At high inputs (50 kg N ha1 yr1) biological N fixation was nearly completely absent. The combined effect of decreased N fixation and the sequestration of N into feather moss tissue accounted for more than half of the N input (56,7%) at the lowest simulated Ndeposition rate. At the highest rate, the feather moss tissue was no longer acting as a sink for N. Rather, a net release of N from bryophyte tissue occurred. Gundale et al. (2011) suggest that this bryophyte effect is partly responsible for why soil inorganic N availability and acquisition by woody plants often remain unchanged at low N addition rates. Vascular plants cannot acquire anthropogenic N inputs until the bryophyte layer is N saturated Effects of WTH We are not aware of any studies investigating the impact of WTH on Nfixation rates. For effects of WTH on bryophytes and lichens, see Appendix D Effects of SH We are not aware of any studies investigating the impact of SH on Nfixation rates. For effects of SH on bryophytes and lichens, see Appendix D. With regard to nonsymbiotic fixation, there are very few studies (Johnson, 2006), but Verburg et al. (2004) found no effect of elevated CO2 on nonsymbiotic fixation of N Effects of N It is well established that Nfixation rates in cyanobacteria are directly influenced by soluble N concentrations (Paul & Clark, 1996) and N additions to a latesuccession forest in the north of Sweden was found to have dramatic and lasting influence on Nfixation rates (Zackrisson et al., 2004; DeLuca et al., 2007). N fertilization, at a rate of 4.5 kg ha1 yr1, basically eliminated any N fixation throughout a threeyear period. DeLuca et al. (2008) also reported a close relationship between increasing N in throughfall and decreasing N fixation, suggesting an ecosystemlevel feedback in which N bioavailability in early successional yields increased N deposition on the moss carpets via canopy throughfall and litter, which downregulates N fixation. Other studies have also shown an inverse relationship between rates of biological N fixation and N availability (see for example Gundale et al., 2011). Recently, Gundale et al. (2011) demonstrated that bryophytes actually attenuate anthropogenic N inputs in boreal forests by limiting the availability of anthropogenic N to vascular plants by directly offsetting N capital as a result of downregulation of N fixation and through sequestration of N capital into their biomass. This effect is most pronounced at low rates of N input, with the lowest N addi13

16 Figure 1. The total deposition of inorganic N (NH 4 and NO 3 ) to coniferous forests in Sweden as modelled with the MATCH model. In order to reduce the effect of extreme years, the average values for 2009, 2010 and 2011 have been used. For more information, see 14

17 3.2 Atmospheric deposition of N Human activities have led to major increases in the global emissions of N to the atmosphere, and, consequently, to an increase in the deposition of biologically available N to the terrestrial biosphere, particularly in Europe and North America. On a global basis, it is estimated that in 1860, 34 Tg N yr 1 was emitted as NOx and NH 3 and then deposited to the Earth s surface as NOy and NHx. In 1995, this value had increased to 100 Tg N yr 1, and by 2050, it is projected to be 200 Tg N yr 1 (Galloway et al., 2008). As a comparison, N deposition to ecosystems in the absence of human influence is generally around 0,5 kg N ha 1 yr 1 or less (Galloway et al., 2008). According to Galloway et al. (2008), there are now large areas of the world where the average Ndeposition rates exceed 10 kg N ha 1 yr 1. By 2050, some regions may reach N deposition levels of 50 kg N ha 1 yr 1 (Galloway et al., 2008). Measuring N deposition on more than 200 sites in 23 countries in the end of the 1990 s, De Vries et al. (2003) showed that the amount of inorganic N deposited in Europe ranged from 1.4 to 42 kg ha 1 yr 1. Approximately 55% of the plots got an external input of N exceeding 14 kg ha 1 yr 1 (De Vries et al., 2003). The measurements by De Vries et al. (2003) confirmed that there is still a clear gradient in the deposition of N compounds, starting in northern Scandinavia (boreal forests), increasing towards southern Scandinavia (borealtemperate forests) and reaching its peak in western, central and eastern Europe. The gradient within Sweden has been confirmed by for example HallgrenLarsson et al. (1995) and PihlKarlsson et al. (2011). In the latter study, the openfield deposition of inorganic N decreased from 11 kg ha 1 yr 1 in the southwest to 2 kg ha 1 yr 1 in the north, reflecting the prevailing wind direction carrying pollutants from the continent to the southwestern part of Sweden (note that dry deposition is not included in these measurements). The relative contribution of NH 4 and NO 3 to European N deposition varied largely over plots, but in general, there was a weak but significant correlation between both N species (De Vries et al., 2003). In the boreal and borealtemperate region, the deposition of NH 4 was generally almost twofold that of NO 3, with values ranging from mol c ha 1 yr 1 for NH 4, and mol c ha 1 yr 1 for NO 3. Most of the variation in deposition could be explained by region, although the deposition of NH 4 was also strongly influenced by the amount of precipitation and altitude (De Vries et al., 2003). average 20% of the total N deposition. The organic N deposition varied between 0.2 and 4.0 kg ha 1 yr 1 with a median value of 1.1 kg ha 1 yr 1. The highest deposition values were found in the southern parts of the country, but there was no clear gradient from the south to the north of the country Throughfall In two ecosystem manipulation studies, covering twelve sites with varying atmospheric N inputs, Tietema & Beier (1995) found a strong correlation between N input by precipitation and the throughfall of both NO 3 and NH 4. In general, their sites, dominated by coniferous forests in central and western Europe, had higher concentrations of NH 4 and NO 3 in throughfall than in bulk precipitation, indicating the importance of dry deposition as well as exchange processes in the canopy for the amount of N reaching the soil. Furthermore, they found that NO 3 dominated sites had higher N concentrations than NH 4 dominated sites, despite the same input by throughfall. According to Tietema & Beier (1995), this implies a higher NH 4 uptake in the canopy in the NO 3 dominated sites (with generally low N input), compared with the NH 4 dominated sites (with generally high N inputs). Also De Vries et al. (2003) found higher NH 4 /NO 3 ratios in bulk deposition (1.5) than in throughfall (1.14), indicating a preferential uptake of NH 4 in the canopy of trees. In contrast to the forest sites investigated in Tietema & Beier (1995), forests in southern and central Sweden have generally shown a considerable canopy uptake of N, resulting in substantially lower concentrations of N in throughfall compared with wet deposition in adjacent fields (Westling et al., 1995). In areas with high deposition of N, like coastal areas in the southwest of Sweden, the dry deposition may be substantial though and N in the throughfall may then exceed that in the wet deposition (Westling et al., 1995). BalsbergPåhlsson & Bergkvist (1995) found, for example, that the supply of N was increased by a factor of 3.4 in a spruce forest front (1 m from the forest edge) in comparison to the bulk deposition in the open field. The flux of N then decreased exponentially with distance from the forest edge. Twentyfive meters into the forest, where the edge effect disappeared, N in the throughfall was 1.5 times higher than that of the bulk deposition. This effect was apparent for both NH 4 N and NO 3 N. An adjacent beech forest gave similar results with Several studies have shown that organic N deposition may contribute substantially to total N deposition (Cornell et al., 1995; 2003; Cape et al., 2001; Neff et al., 2002; Ham & Tamiya, 2006). According to Neff et al. (2002), approximately 23% of the N deposited in Europe constitutes organic N. Most likely, this proportion has not changed much during the 20 th century (Cape et al., 2001). The contribution of organic N to total N deposition in Sweden seems to be similar to the values estimated for Europe in general. Measurements on 24 open field sites within the SWETHRO network (PihlKarlsson et al., 2011) during October 2011 to September 2012 showed that organic N constituted on 15

18 regard to the throughfall. However, in this stand, the edge effect was not pronounced (BalsbergPåhlsson & Bergkvist, 1995). The filtering effect of tree canopies is generally obvious in all forest types, although different species may have different filter capacity, with coniferous forests usually being more effective in capturing atmospheric particles than deciduous forests (BalsbergPåhlsson & Bergkvist, 1995). Bergkvist & Folkesson (1995) found that the deposition of NH 4 N and NO 3 N was 1.5 to 3 times higher in a spruce stand than in two deciduous forest stands (with deposited amounts to the two deciduous stands, consisting of one beech and one birch stand, being quite similar). The factors determining the filter capacity of canopies was suggested by BalsbergPåhlsson & Bergkvist (1995) to be total leaf area (particularly different when deciduous trees are defoliated during winter) and leaf surface characteristics together with the general structure of the trees. 3.3 N fertilization Fertilization with N has been used as a measure to increase tree growth since the mid 1960 s. In Sweden, around 2 million ha (or 10% of the productive forest land) has been subjected to N fertilization on some occasion (Nohrstedt, 2001). When at its highest peak, in the late 1970s, nearly ha were fertilized annually. In the end of the 90 s, this number had decreased to ha (Nohrstedt, 2001). Between approximately 1995 and 2005, the area that was fertilized varied between and ha (Swedish Forest Agency, 2007). In 2009, this number had increased to ha (Swedish Forest Agency, 2010). Both urea, NH 4 NO 3 and NH 4 NO 3 amended with dolomite have been used as N fertilizers in Sweden. Nowadays, the recommendation from the Swedish Forest Agency is to apply maximum 200 kg N ha 1 of NH 4 NO 3 amended with dolomite, with an intermediate interval of at least 8 years (Swedish Forest Agency, 2007). To avoid negative effects, the total amount that should be added during a rotation period varies depending on the geographical location. The recommendations for the different parts of the country can be seen in Figure 2. Figure 2. The Swedish Forest Agency s recommendations for N fertilization in Sweden. In the southernmost part of the country (dark green and middle green), no fertilization is allowed (an exception is if wholetree harvesting is applied in the middle green area). In the middle and the northern part of the country (light green and yellow areas), fertilization is allowed but should not exceed 300 kg (light green) and 450 kg (yellow area) of N per hectare and forest generation. For a more detailed description, see Swedish Forest Agency (2007). Previously, both experimental and commercial rates of fertilizer addition have often exceeded the recommended amounts, and application rates of several hundred kilos of N per hectare have been reported in Sweden as well as in other countries (Aber et al., 1989; Andersson et al., 1995; 1998; Tamm et al., 1999; Jacobson & Pettersson, 2001). In many of these studies, fertilizer was applied annually or with only a few years interval. The rates of atmospheric deposition of N are generally low compared with the rates of fertilizer application. Furthermore, fertilizer applications are often one or a fewtime applications, whereas atmospheric deposition is chronic. In this review, N deposition and N fertilization are both considered as input processes, only with different rates. We assume that the resulting effects of N on ecosystem processes depend on the concentrations and forms of N rather than on the source of input. 16

19 4. NITROGEN IN PLANTS 4.1 N acquisition and uptake Approximately 1.5 % of the dry weight of plants is accounted for by atoms of N (Kimmins, 1997). Plants adsorb N in three distinct forms: NO 3, NH 4 and amino acids. The rate of N uptake depends on both the concentration in the environment and the demand of the plant, the latter being at least partly determined by the growth rate of the plant (Lambers et al., 1998). The majority of terrestrial plants absorb N primarily via their roots from the soil (Lambers et al., 1998). However, leaves are also capable of N uptake (Sutton et al., 1995). It is wellknown that plants have several different mechanisms to enhance the availability of nutrients in the rhizosphere. With regard to N, mycorrhizal associations are regarded as one of the most important (Smith & Read, 2008). Roots of boreal forest trees are usually heavily colonized by ectotrophic fungi. In Norway spruce, for example, more than 90% of the root apical zones are usually enclosed by a fungal sheath (Marschner, 2003), and microcosm studies of EM fungi have shown that 20 to 30 % of the C assimilated by the host plant may be consumed by the fungal partner (reviewed in Söderström, 2002) Effects of temperature According to Lambers et al. (1998), low temperature directly reduces N uptake by plants, as expected for any physiological process that is dependent on respiratory energy. Rennenberg et al. (2009) emphasized the temperature sensitivity of the highaffinity transport system (HATS) prevalently used in forest ecosystems with low soil solution concentrations of N. For NO 3 uptake, a consistent increase of the HATS activity was observed when root temperature increased in a range from approximately 5 to 25 C. The strong response of HATS to root and soil temperature may, according to Rennenberg et al. (2009), be explained by the enzymatic nature of the HATS (as proposed by Glass et al., 1990). Several other studies also suggest that nutrient uptake by trees increase with rising temperatures, at least until a certain treshold value (Bassirirad, 2000). For spruce and beech, Geßler et al. (1998) found a maximum uptake of NH 4 at 20 C and 25 C, respectively, followed by a decrease at higher temperatures. Maximum uptake for NO 3 was found to be at 25 C. According to Lukac et al. (2010), such high temperatures are not likely to be reached under field conditions, at least not in a closed canopy forest. Geßler et al. (1998; 2005) observed soil temperatures ranging from 10 to 16 C in the upper soil layers of temperate beech and spruce forests during the growing season. At temperatures of 10 and 15 C, net uptake of NO 3 by spruce and beech amounted to around 16% and 11% of maximum uptake at 25 C. At 10 C, net uptake of NH 4 reached 73% and 31% of maximum uptake for spruce and beech (Geßler et al., 1998). However, in open canopies (such as after storm events, clearcutting or intensive selective felling), soil temperatures in the uppermost soil layer may approach the optimum for net Nuptake (Rennenberg et al., 2009). That soil temperatures affect NO 3 and NH 4 uptake of trees has also been observed in the field (Geßler et al., 1998). For spruce and beech growing in temperate forests, a seasonal pattern with maximum uptake rates of NH 4 in midsummer and strongly restricted uptake rates in spring and autumn when soil temperatures were below approximately 12 C was found (Geßler et al., 1998). According to Rennenberg et al. (2009), N uptake from the soil also depends on transpiration rate. It thus remains to be elucidated whether higher soil temperatures in spring increase early season N uptake of roots since transpiration is either low (evergreens), or not present (deciduous). Temperature also seems to influence the proportions of NH 4 and NO 3 taken up by trees (Rennenberg et al., 2009). At lower soil temperatures (up to 10 C), NO 3 uptake is minimal, whereas NH 4 uptake already reaches >50% of its temperaturedependent maximum in spruce and beech. With increasing soil temperature (>10 C), the preference for NH 4 is generally maintained but the relative proportion of NO 3 increases (Geßler et al., 1998) Effects of moisture Reduced soil water availability and extended drought periods are known to reduce fineroot biomass (Cudlin et al., 2007; Rennenberg et al., 2009), thus decreasing the nutrient absorbing surface of root systems. Both drought and excess water availability may also impair the mineral nutrition of trees by influencing: 1) the nutrient availability in the soil and 2) the physiology of the uptake systems of the mycorrhizal tree roots (Kreuzwieser & Geßler (2010). According to Rennenberg et al. (2009), not much is known about the effect of increased drought on the physiology of N transport. Most studies to date allow the characterization of the N status of the tree, but do not provide information as to whether changes in the N status are due to changes in soil water availability or root Nuptake capacity, or both. Applying 15 N labelling, Geßler et al. (2005) assessed NO 3 and NH 4 uptake kinetics of nonexcised intact mycorrhizal roots of adult beech trees on two aspects of a site that differed in soil water availability. At the drier aspect, the maximum rate of NO 3 uptake (and thus uptake capacity) was reduced by more than 50% as compared with the wetter aspect during most of the growing season. As differences in NO 3 availability, rooting patterns and in the affinity of the transport system were ruled out as responsible for the difference, a longterm effect on NO 3 transport of continued water depletion was assumed. Whether differences in 17

20 NO 3 transporter abundance/activity, the expression of different NO 3 transporters and/or mycorrhizal colonization are responsible for the droughtdriven reduction remains to be clarified (Rennenberg et al., 2009). Shi et al. (2002) observed large changes in the species composition of fungal partners in beech mycorrhiza as a consequence of drought and also an accumulation of stressrelated metabolites in the mycelia, indicating that mycorrhizal fungi may exert a strong influence on the drought reaction of plants. The droughtsensitivity of N acquisition in beech was also emphasized by Geßler et al. (2004) and Geßler et al. (2002) found that net uptake of mycorrhizal roots of beech was closely related to the transpiration rate of trees Effects of CO 2 Increased levels of CO 2 generally result in an increased allocation of C to belowground, with subsequent increases in the growth of roots and mycorrhiza (see Appendix A, section ), which should allow the trees to exploit a larger soil volume and thus might be expected to promote NO 3 and NH 4 uptake (Stitt & Krapp, 1999). In several of the FACE experiments, the increase in biomass production was indeed accompanied by an increase in N uptake by trees, resulting in a negligible variation in Nuse efficiency (NUE) between trees subjected to ambient and elevated CO 2 (Finzi et al., 2007). Johnson (2006), reviewing the implications for longterm responses of forests to elevated CO 2, concluded that, more often than not, N uptake is increased with elevated CO 2 and the increase is a function of increased root growth, since most studies to date have shown little effect, or none at all, of elevated CO 2 on soil Nmineralization rates (see section 5.1.3). However, the relationship between elevated CO 2 and N uptake is not completely straightforward (see for example review by Bassirirad, 2000). Calfapietra et al. (2007), investigating the effect of elevated CO 2 on a shortrotation poplar plantation, found no change in cumulative N uptake by trees over the rotation as a consequence of elevated CO 2, despite a substantial increase in biomass production. This resulted in a significant increase in NUE and a decrease in the concentration of N in most tissues. Several examples of increases in NUE are also given in the review by Stitt & Krapp (1999). The authors suggested that the decreased N concentrations observed under elevated CO 2 indicate that N uptake and assimilation often fail to keep pace with photosynthesis and growth under elevated CO 2. However, they emphasized that there is no hinderance for elevated CO 2 to lead to increased N uptake, provided there is an adequate supply of N outside the plant. Bassirirad (2000), on the other hand, suggested that differences in 18

21 the experimental protocols as well as differences in speciesspecific responses were the major determinants behind the variable results. Alberton et al. (2007) found that mycorrhizal growth (extraradical hyphal length) under elevated CO 2 was negatively correlated with shoot Ncontent and total plant Nuptake. According to Lukac et al. (2010), one explanation for this seemingly contradictory observation of N uptake may be the observed increase in fungal abundance and proliferation and the increase in fungi/ bacteria ratio in soils under elevated CO 2 (Treseder, 2004; Hu et al., 2006; Parrent et al., 2006; Carney et al., 2007). Fungi have higher C/N ratios than soil bacteria, thus using smaller amounts of N to produce equivalent amounts of biomass. Furthermore, they translocate C and N within the fungal mycelium (Boberg et al., 2010), something that according to Lukac et al. (2010) may explain the low mineralization rates and, hence, the lower N availability in fungal dominated ecosystems. Lukac et al. (2010) concluded that the positive effects of elevated CO 2 on infection rates of mycorrhiza (Hu et al., 2006), mycorrhizal activity (Lukac et al., 2009) and turnover (Godbold et al., 2006) may enhance tree nutrition in the future, but only if mycorrhizal fungi proliferate at the expense of bacteria or other functional types. In one of the few studies examining the impact of elevated CO 2 on organic N uptake, no effect of elevated CO 2 was reported (Hofmockel et al., 2007) Effects of N While increasing N supply usually enhances both shoot and root growth, shoot growth is generally more stimulated than the root growth, leading to a fall in rootshoot dry weight ratio with subsequent implications for nutrient uptake in the long term (see appendix A, section 4.3.4). However, the effectiveness of N inputs seems to be limited by immobilization and other mechanisms. For example, when labelled N was added to soil and litter in a forest over seven years, only a small fraction became available for tree growth (Nadelhoffer et al., 2004). Atmospheric deposition of N to the surface of leaves may reduce N uptake from the soil. The flux of N from soil into the roots has been shown to be downregulated to an extent that equals the N influx into the leaves (Rennenberg & Geßler, 1999) Effects of WTH There is basically no scientifically published data on how removal of slash may affect N uptake and acquisition of trees in the subsequent generation. One exception is the study by Weatherall et al. (2006a). By labelling growing Sitka spruce (Picea sitchensis (Bong.) Carr.) seedlings with 15 N, then harvesting the aboveground biomass and placing it on soil in a pot containing newly planted seedlings, they showed that N in brash can make a small, but discernible, direct contribution to new tree growth one growing season after the application of the brash. The authors emphasized, however, that the contribution is likely to vary considerably depending on site conditions, and that field studies are necessary before conclusions can be drawn about the effects of slash removal on N uptake. With regard to the influence of WTH on N content in litter and forest soils, see sections and Effects of SH Using a similar method as above (stable isotope technique; Weatherall et al., 2006a), Weatherall et al. (2006b) traced nutrient release from decomposing roots and the subsequent uptake into newly planted Sitka spruce seedlings. They found that decomposing roots contributed up to 310% of the N subsequently taken up by new trees after one single growth season. The percentage contribution from decomposing roots to new tree growth was markedly higher in a nutrientpoor soil than in a more nutrientrich soil. With regard to the influence of SH on N content in litter and forest soils, see sections and N content of trees, allocation, retranslocation and storage According to Lukac et al. (2010), total N uptake and N allocation of trees are two of the main factors controlling foliar N concentration. In theory, trees optimize their N allocation for attaining maximum growth, thus allocating available N to organs with greatest benefit for net growth. It is wellestablished that trees store N, and that the remobilization of stored resources can provide an appreciable proportion of the annual N requirement used for aboveground growth each year (Millard et al., 2007). Deciduous plantation trees of oak and larch in Wisconsin, USA, were found to have retranslocation values for N of around 80%, supplying on average 77% of the aboveground annual N requirement (Son & Gower, 1991). Evergreen plantation trees in the same study retranslocated between 13 and 55% of their N from senescing foliage, supplying on average 27% of the aboveground annual N requirement (Son & Gower, 1991). For Scots pine in Scandinavia, values have been found to be even higher, approaching 70 and 80% in some studies (Helmisaari, 1992a; Nieminen & Helmisaari, 1996), thus supplying between 30 and 50%s of the N required for annual biomass production (Helmisaari, 1992b). Norway spruce seem to have lower retranslocation efficiencies than pine (Bothwell et al., 2001), with values found in the literature ranging from 13 to 36% (Son & Gower, 1991; Bothwell et al., 2001). The information with regard to the effects of climate change and forest management on storage and retranslocation of N in trees is generally sparse. Most of the information below thus refers to the effects of climate change and forest management on N concentrations in trees Effects of temperature According to a review by Pendall et al. (2004), investigating the responses of belowground processes to elevated temperatures and CO 2, elevated temperatures are generally reported to increase the root Nconcentration, probably because mineralization and diffusion of N are stimulated at high temperatures. In coherence with the results of Pendall et al. (2004), Tingey et al. (2003) found increased leaf N concentrations in Douglas fir as a consequence of elevated temperature. However, Lukac et al. (2010) suggested 19

22 that the nature of the temperature increase is of importance for its impact on the nutrient content. While a small and gradual increase in mean annual temperature may give the above mentioned results, a period of very high seasonal maximum is likely to have the opposite effect, not least because of its interaction with drought. In accordance with the statement of Lukac et al. (2010), Fotelli et al. (2005) found that elevated temperature and high irradiance resulted in impaired N acquisition of drought sensitive beech seedlings (Fagus sylvatica) Effects of moisture Drought generally causes an impairment of the nutrient contents and concentrations in trees (Kreuzwieser & Geßler, 2010). Already moderately lower water availability may result in reduced soluble N contents, as was shown by Nahm et al. (2006) for various tissues of European beech. Also Ge et al. (2010), investigating the impacts of climate change on the productivity of Norway spruce dominated mixed stands in relation to water availability in southern and northern Finland using a processbased ecosystem model (FinnFor), suggested that increasing soil water deficit under a changing climate will result in a reduced N uptake of trees in Finland. Over the 100year simulation period, the changing climate resulted in a clearly reduced moisture content of the SOM layer. Despite this, the gross N content was fairly stable due to the continuous accumulation of litter and plant debris on the forest floor. However, the amount of decomposed soil organic matter (humus as available source of N) decreased during the latter phases of the simulation period. Consequently, a changing climate with more frequent drought episodes resulted in a decrease in the N uptake, and subsequently also lower N content in the needles of trees. These results are logical considering the reduced inputs of labile C to the soil under drought stress (see for example Ruehr et al., 2009), implying that the supply of energy and C skeletons for plant nutrient uptake and assimilation is likely to be constrained. Furthermore, Fotelli et al. (2002) noted a reduced uptake of 15 N when beech seedlings were grown under simulated drought conditions and in competition with Rubus fruticosus in a greenhouse study. Due to protein degradation, there was an increase in soluble nonprotein N in roots and leaves of the seedlings. The protein degradation was attributed to increased levels of amino acids serving as osmoprotectants under drought conditions Effects of CO 2 The initial stimulation of photosynthesis under elevated CO 2 is often accompanied by a decrease in N concentration of the plant (Stitt & Krapp, 1999). In particular, a decrease in foliar N has often been reported (Tingey et al., 2003; Ellsworth et al., 2004). This phenomenon seems to be particularly marked when N supply is low, whereas when N supply is adequate there is no acclimation of photosynthesis, no major decrease in internal N concentration or the levels of N metabolites (Stitt & Krapp, 1999). According to Lukac et al. (2010), the decrease in plant N might be a result of either decreased N uptake per unit of biomass produced, or preferential allocation of acquired N to other tissues, i.e. fine roots (which has not been much studied). However, according to a review on the effects of elevated CO 2 on belowground compartments by Pendall et al. (2004), elevated CO 2 is expected to also result in decreased fineroot concentrations of N (by 1025%). The process of N retranslocation from leaves of various poplar species (before abscission) has been found to be slightly increased by elevated CO 2. This was accompanied by more N being immobilized in woody tissues (Calfapietra et al., 2007). Since woody tissues have substantially longer turnover times, Lukac et al. (2010) thus suggested that N retranslocation could enhance N immobilization in trees Effects of N Mellert et al. (2004), investigating longterm nutritional trends of 49 conifer stands in Scandinavia and Central Europe, found that atmospheric deposition of N has resulted in increased concentrations of foliage N in Scots pine grown predominantely on naturally poor or devastated soils. For Norway spruce, on the other hand, the results were the opposite, i.e. a negative trend with regard to foliar N concentrations. However, since there was an overall increase in the foliar N content, Mellert et al. (2004) concluded that the decline was most likely the result of dilution. Fertilisation experiments with N have generally shown increased leaf N concentrations (Nohrstedt, 2001 and references therein), although the increases may sometimes last only for a relatively short period of time (Mälkönen & Kukkola, 1991). When plants are supplied with rather high levels of NO 3 from the soil, N is generally stored in the form of NO 3. At a moderate or low N availability, on the other hand, N is stored as amino acids (often of a kind not found in proteins), amides (asparagine and glutamine), or proteins (Lambers et al., 1998). There is evidence that in a range of both deciduous and coniferous evergreen trees, a significant proportion of the N that is subsequently used for leaf growth is stored as Rubisco (Millard et al., 2007). However, it may also be stored in special storage proteins. Storage as protein involves the additional cost of protein synthesis, but has no effect on the cell s osmotic potential. In addition, proteins may serve a catalytic or structural function, as well as being a store of N (Lambers et al., 1998). The fraction of mineral nutrients withdrawn from leaves before their abscission is variable with differences associated with leaf area span, overall leaf nutrient content, soil fertility, and environmental conditions such as drought (Niinemets & Tamm 2005; Del Arco et al., 1991; Kimmins, 1997). In a review comparing over a hundred deciduous and evergreen shrubs and trees, Aerts (1996) reported that 50% of the N was withdrawn from the leaves. In 32% of the experiments analysed, there was a decrease in N resorption in response to increased nutrient availability, while in 63% of the cases there was no response. Other studies have also indicated that the efficiency of N withdrawal and retranslocation are in many cases unaffected by soil fertility (Chapin & Moilanen, 1991; Helmisaari, 1992b), but are controlled by internal factors, primarily growth rate (Nambiar & Fife, 1987; 1991). Ladanai et al. (2007) showed that the response of boreal forest ecosystems to longterm NPK additions is timedependent and modifies 20

23 21

24 N pools in a speciesspecific and sitespecific way. When comparing a Norway spruce stand with a Scots pine stand in Central Sweden, they found that while the soil was the major sink for N in the pine stand, the trees were the major sink for N in the spruce stand. In the spruce stand, the increase of N in the soil was restricted to the humus layer Effects of WTH Olsson et al. (2000) reported lower concentrations of N in currentyear needles of Scots pine and Norway spruce stands after WTH compared with CH. The difference occurred 810 years after clearcut, but had diminished when the stand reached the age of 1618 years. Luiro et al. (2010), investigating a series of 12 longterm (3 to 20 years) experiments in Finland, found that N was sometimes (in one Scots pine and one Norway spruce stand) significantly lower after WTH compared with CH after the second thinning. However, the differences later disappeared, and none of the trees in the experiments showed a significantly reduced growth as a consequence of lower concentrations of nutrients. In coherence with the studies mentioned above, most longterm studies (>20 years) have shown no effect of WTH on the N concentrations of trees. Saarsalmi et al. (2010), for example, found no significant effects on needle concentrations of nutrients as a consequence of harvest intensity in two Scots pine stands in eastern Finland 22 years after clearcut. One stand was growing in a fertile soil, while the other was growing in a less fertile soil. Wall & Hytönen (2011) also found no significant effects on concentration of foliar nutrients when comparing WTH with CH 30 years after clearcut in a Norway spruce stand in central Finland. However, in the latter study, needles were left on site at WTH, supplying the trees with a majority of the nutrients otherwise removed at WTH. 4.3 Plant growth and aboveground production For a description of plant growth and aboveground production and how the various factors influence them, see Appendix A section Root growth and mycorrhiza production For a description of root growth and mycorrhiza production and how the various factors influence them, see Appendix A, section Litterfall The N that is not recycled within the tree returns to the soil as litter. The quantity of N transferred from trees to soil by aboveground litterfall is a function of the biomass, the type (i.e. leaves, branches, bark etcetera), and the nutrient concentration of the litterfall, all of which vary from site to site. Litter production in Sweden is clearly related to the aboveground tree biomass (see for example Berggren Kleja et al., 2008) and, accordingly, litter production here show a clear southnorth gradient. A review by Kimmins (1997) showed that the quantity of N in aboveground litterfall in various forests generally varied between 10 and 70 kg of N ha 1 yr 1 in the northern hemisphere. In many types of forest, shrub and herbaceous communities, there is a greater annual turnover of organic matter belowground than aboveground. Although very little studied, recent investigations have shown that belowground litterfall is a major pathway of nutrient loss from plants (Kimmins, 1997), and it does seem like fineroot litterfall varies according to both stand age and site productivity (Vogt et al., 1983; Attiwill & Adams, 1993) Effects of SH In mature forest stands, stumps account for around 815% of the N found in tree biomass (Palviainen et al., 2010 and references therein). To our knowledge, only two studies have evaluated the effects of SH on foliar chemistry of trees planted after harvest. Hope (2007) investigated the effects of SH in combination with soil scarification on foliar chemistry of lodgepole pine (Pinus contorta) and hybrid spruce (Picea glauca (Moench) Voss x Picea engelmanii Parry) at three sites in Canada. He found that foliar levels of N in spruce were significantly lower where stumps had been harvested. For pine, there was no significant difference. In five Douglas fir (Pseudotsuga menziesii (Mirb.) Fanco) stands in the US, trees growing in plots where SH had been applied 2229 years prior to sampling also showed lower foliar N concentrations than the trees growing in plots where no stumps had been removed (Zabowski et al., 2008). The difference was not significant, but since it had persisted for more than 20 years Zabowski et al. (2008) suggested that additions of N or organic matter should be applied to ameliorate any longterm losses of soil N when stumps are harvested. In the study by Zabowski et al. (2008), all sites had been fertilized with various amounts of N before planting Effects of temperature Higher temperatures are generally associated with higher NPP (see Appendix A section 4.2.1). Provided no other resources are limiting, increased temperature thus has the potential to provide more substrate to the soil (Pendall et al., 2004) Effects of moisture In general, both total aboveground litterfall and leaf litterfall of forests increase from polar regions towards the equator (Kimmins, 1997). This parallels the increase in biomass and NPP, since aboveground litterfall generally reflects aboveground productivity. Consequently, litterfall losses are generally greatest on moist, warm, fertile and other highproductivity sites and least on dry, cold, infertile and other lowproductivity sites (Kimmins, 1997) Effects of CO 2 The higher production under elevated CO 2 (see Appendix A section 4.2.3) will most likely produce additional litter (Schlesinger & Lichter, 2001), at least if the N availability is high (Zak et al., 2000). According to the review by Hyvönen et al. (2007), elevated 22

25 CO2 is also expected to produce energyrich but nutrient poor litter with higher C/N ratios. Norby et al. (2001) found that N was reduced by on average 7.1% and lignin increased by 6.5% in leaves of plants grown at elevated CO2 compared with those grown at ambient levels. Other studies have also showed effects of elevated CO2 on leaf chemistry, for example decreased N concentrations (Norby et al., 1999) and changes in foliar concentrations of starch (Kainulainen et al., 1998). Chapin et al. (2009), linking plantsoil C dynamics to global consequences, emphasized that elevated CO2 can give rise to litter that is more resistant to microbial breakdown. In contrast to the studies mentioned above, Verburg et al. (1999) found decreased lignin concentrations in leaves of Betula pubescens as a consequence of elevated CO Effects of N An increase in the amount of available N has in many studies been associated with increased growth and increased leaf concentrations of N (see for example Nohrstedt, 2001), and litter production in Sweden has been shown to be related to the aboveground tree biomass (see for example Berggren Kleja et al., 2008). Subsequently, addition of N may be expected to result in increased amounts of litterfall, and litterfall with higher N concentrations. A decrease in the rate of root turnover with increasing litterfall and return of N from plant to soil was reported by Attiwill & Adams (1993) for broadleaved forests across the world. Their results were based on results by Vogt et al. (1986) as well as on their own data. The authors suggested the decrease to be a function of the decrease in root:shoot ratio with increasing productivity Effects of WTH In a forest subjected to WTH, nutrients contained in the slash are removed from the system, with the result that nutrients are permanently lost from the ecosystem. Akselsson & Westling (2005) have compiled data on N concentrations in stems, branches and needles of various tree species and used them for calculations of N losses at different harvest rates and scenarios. They found that the average concentration of N in stems were 1.1 mg g1 for spruce and 0.9 mg g1 for pine. The corresponding values in branches were 6.0 and 3.4 mg g1 and in needles 11.2 and 12.4 mg g1. Similar values were found by Alriksson & Eriksson (1998; 1.8 and 1.5 mg g1 in stems, 5.1 and 4.3 mg g1 in branches and 11 and 15 mg g1 in needles of spruce and pine, respectively) and by Thelin (1998; and 1415 mg g1 in needles of spruce and pine, respectively). amount in spruce stands was approximately twice that of pine, i.e kg N ha1.the impact of slash removal on the composition and chemical quality of the remaining litter has to our knowledge not been investigated Effects of SH In a forest subjected to SH, nutrients contained in the stumps are removed from the system, with the result that nutrients are permanently lost from the ecosystem. The root systems of trees constitute a significant portion of the trees. Walmsley & Godbold (2010) reported that stumps constituted on average 3641% of the total tree biomass for coniferous trees grown in the UK, depending on species. Of the total belowground biomass, coarse and structural roots make up by far the greatest portion, and for Sitka spruce, they were found to constitute 88% (Walmsley & Godbold, 2010). According to Hellsten et al. (2010), N losses at SH are smaller than when for example slash is removed, since N concentrations in stumps are lower than those in slash. Hellsten et al. (2010) emphasized, however, that since the biomass of stumps is substantial N losses are not negligible. Palviainen et al. (2010) suggested the N pool in stumps to be rather large, and estimated that pine stumps contain 1114 kg N ha1, while spruce stumps contain 79 kg N ha1. Walmsley & Godbold (2010) emphasized that, analogous to stems, the amounts of nutrients removed from forest ecosystems at SH depend upon the relative amount of bark to stemwood. They referred to studies suggesting that smaller stumps (less than 15 cm diameter), which consist of a higher proportion of bark and hence nutrients, should not be harvested as their removal is inefficient and uneconomic due to their small root biomass. The impact of the removal of stumps on the composition and chemical quality of the remaining litter has to our knowledge not been investigated. From these data, it is obvious that the amount of N lost through harvesting depends not only on the amount of harvested biomass, but also on what fraction of the tree that is harvested and on tree species. The calculations by Akselsson & Westling (2005) showed that WTH resulted in a two times higher N loss than CH. Furthermore, the difference in N loss between CH and WTH was greater for spruce forests compared with pine forests, since spruce trees have a larger relative fraction of branches than pine trees (Akselsson & Westling, 2005). Based on a review by a number of studies, Palviainen et al. (2010) estimated the pool of N in aboveground logging residues to be kg N ha1 in pine stands. The 23

26 24

27 5. NITROGEN IN SOILS 5.1 Decomposition, mineralization and immobilization When organic tissue is added to an aerobic soil, three general reactions take place: 1. Carbon compounds are enzymatically oxidised to produce carbon dioxide, water, energy and decomposer biomass. 2. The essential nutrient elements, such as N, P and S are mineralized and/or immobilized by a series of specific reactions that are unique for each element. 3. Compounds very resistant to microbial action are formed, either through modification of compounds in the original tissue or by microbial synthesis. Mineralization is a primarily biological process that converts organic N in decomposing litter to inorganic N in the form of NH 4 (Paul & Clark, 1996). It can provide up to 75% of the mineral N input into a forest soil (Berendse et al., 1989). As for decomposition in general, ph, moisture, temperature, and, in particular, soil C and N concentrations and the C/N ratio of the litter material, seem to have the strongest influence on the mineralization rates (Attiwill & Adams, 1993; Paul & Clark, 1996; Kimmins, 1997; Andersson et al., 2002; Booth et al., 2005). Once NH 4 is formed, there are a number of possible fates. It can be taken up by plants and is often a preferred N source in the soil solution. However, it may also be immobilized, i.e. used for microbial growth, held on exchange complexes or enter the interlayer portion of clays. Immobilization can thus be both biotic and abiotic. Bengtsson et al. (2003) reported the extent of immobilization to vary between 35 and 95% depending on soil type and environmental conditions and Brooks et al. (1985) reported that soil microorganisms are responsible for 10 to 50% of NH 4 immobilization (most studies indicate that microbes prefer NH 4 over NO 3, see below). The average turnover time of N in microbial biomass is estimated to between one and two months (Davidson et al., 1992). However, some of the immobilized N becomes very stable as a result of repeated cycles of immobilization and mobilization (He et al., 1988) and accumulation of persistent residues of microbial cells (Paul & Juma, 1981). The influence of climate and forest management on the decomposition process is described thoroughly in Appendix A, section 5.1. The following section thus focuses on their effects on the mineralization and immobilization processes Effects of temperature According to a recent review by Rennenberg et al. (2009), rates of decomposition, N mineralization and nitrification generally increase with temperature. In a metaanalysis of 32 sites (including high and low tundra and forest), Rustad et al. (2001) found a significant increase in net Nmineralization at higher temperatures (on average 46%) for the organic soil horizons at the 12 sites for which data was available. Several other soil warming studies have also found increased net Nmineralization rates as a consequence of increased temperature (Peterjohn et al., 1994; Van Cleve et al., 1990; Shaw & Harte, 2001; Melillo et al., 2002). According to Rennenberg et al. (2009), the temperatureinduced increase in net Nmineralization may be caused by a proportionally larger temperature response of gross Nmineralization than immobilization, which could be related to the rapid assimilation of labile C at elevated incubation temperatures something that may limit the C supply to heterotrophic inorganic Nimmobilizing microorganisms. In contrast to the studies mentioned above, Beier et al. (2008) reported that N mineralization in shrublands across a climatic gradient from Denmark to Spain was relatively insensitive to temperature increases. Instead, it was mainly affected by changes in soil moisture. However, since enzyme activities are temperaturesensitive processes (generally following the Q 10 relationship), and studies have reported higher enzyme activities related to N cycling in soil in response to warming and in the absence of moisture limitations, Lukac et al. (2010) concluded that an increase in N mineralization in soil is to be expected under favourable moisture conditions and substrate availability in those ecosystems where temperature is a limiting factor. Another significant effect of temperature on mineralization may be through future increases in the frequency of freezingthawing cycles (Ollivier et al., 2011). Significant increases in N mineralization in frozen compared with nonfrozen soils representing various soil types and landuse was shown in DeLuca et al. (1992). For soils under natural vegetation, results are contradictory. Matzner & Borken (2008) concluded in their review on whether freezethaw events enhance C and N losses from soils that although freezethaw events seem to increase net mineralization of N in arable soils after thawing, there is no convincing evidence for a general increase in N mineralization in soils under natural vegetation. Furthermore, the increase, if observed, seems to be rather small in relation to the expected annual mineralization rates (Matzner & Borken, 2008) Effects of moisture The effects of drought on the process rates and the sensitivity of mineralization to drought stress in various forest ecosystems are still poorly understood (Rennenberg et al., 2009) and results are variable. Geßler et al. (2005), for example, found increased gross ammonification rates in summer on a droughtexposed beech site as compared with a coolmoist site, while Johnson et al. (2002; using NO 3 and NH 4 concentrations in soil solution as a proxy for 25

28 plant or ecosystem N availability) and Beier et al. (2008) reported that net Nmineralisation was strongly inhibited by reduced soil water availability. However, in general, reduced soil water availability limits microbial activity in soils, and might thus (depending on the intensity and duration of the drought event) lead to total inhibition of microbial activity (Paul & Clark, 1996; Borken & Matzner, 2009). Rennenberg et al. (2009) emphasized the paradox that despite a decrease in microbial activity as a consequence of drought stress, the availability of DON may increase due to the dieback of microbial biomass, thereby promoting plant N uptake (Borken & Matzner, 2009). However, according to Kreuzwieser & Geßler (2010), it is not yet clear whether plants may actually benefit from this potential N source under drought conditions, since diffusional limitations or other constraints might prevent N uptake and utilization. Drying and rewetting of soils have been shown to result in a N mineralization flush, which may be promoted by accumulated plant and microbial necromass (Borken & Matzner, 2009). However, according to the review by Borken & Matzner (2009), the cumulative net Nmineralization of dryingrewetting treatments compared with permanently moist controls is commonly lower, since the shortlived wetting pulses cannot compensate for the low mineralization rates during the drought periods. An increasing frequency and intensity of dryingrewetting cycles, as expected under climate change, can thus, according to Borken & Matzner (2009), be expected to reduce overall mineral N availability in the soil Effects of CO 2 The results with regard to the effects of elevated CO 2 on mineralization are inconclusive. According to the reviews by Johnson (2006) and Lukac et al. (2010), N mineralization has been found to increase (see for example Zak et al., 1993), decrease (see for example Billings & Ziegler, 2005) or not change (see for example Zak et al., 2003; Holmes et al., 2003; Hoosbeek et al., 2006). Johnson (2006) described the results as perplexing, in that CO 2 often stimulates additional N uptake and even soil N depletion under elevated CO 2 without any measurable change in soil N mineralization or soil N availability. One possible explanation is, according to Johnson (2006), that the measures of soil N mineralization and soil N availability are not sensitive enough to detect changes in mineralization rates. Another explanation may be that the increased N uptake is due to increased root growth and thus soil exploration (Johnson, 2006; see also section 4.4). Both Lukac et al. (2010) and Johnson (2006) concluded that most studies to date have indicated little or no effect of elevated CO 2 on soil Nmineralization rates. According to Singh et al. (2010), it is generally accepted that increased levels of CO 2 quantitatively and qualitatively alter the release of labile sugars, organic acids and amino acids from plant roots, which may stimulate microbial growth and activity. In the long term, it is argued that this increase in microbial biomass can lead to immobilization of soil N, consequently limiting the N available for plant uptake and creating a negative feedback that constrains future increases in plant growth. This may lead to a higher C/N ratios, something that might favour fungal dominance and diversity, and subsequently lower respiration rates and increase the potential for C sequestration. Increased microbial immobilization as a consequence of elevated CO 2 has indeed been reported in the literature (FACE experiments; Zak et al., 2000). However, there are also studies showing no effect at all on immobilization (Zak et al., 2003). Furthermore, several studies have shown that increased levels of CO 2 lead to substantial increases in soil respiration (see review by Singh et al., 2010), further complicating the reasoning presented by Singh et al. (2010). According to Lukac et al. (2010), the outcome depends on the prevalence of N or C limitation in microbial communities and on the C/N ratio of the substrate being decomposed. When decomposing substrates with high C/N ratio, microorganisms will retain more inorganic N (mainly as NH 4 ) during decomposition, thus reducing the availability of this N pool to plants. Conversely, if the C/N ratio of the substrate is lower than that of the decomposers, microorganisms will increase the size of the mineralized N pool in the soil. In the CLIMEXproject in Norway, both temperature and CO 2 were manipulated to simulate the effects of climate change on a boreal forest ecosystem consisting of a mixture of pine (Pinus sylvestris) and birch (Betula pubescens). Within three years, soil N mineralization had increased significantly, according to the authors most likely due to the elevated soil temperatures (Van Breemen et al., 1998) Effects of N The importance of N availability for mineralization has been demonstrated in several studies. A negative correlation between gross Nmineralization and the soil C/N ratio was found by Booth et al. (2005), when reviewing the controls on N cycling in terrestrial ecosystems worldwide. The importance of the chemical quality of the litter was further emphasized by Manzoni et al. (2008). In their study, based on 2800 observations across the world, the critical N/C ratio below which net immobilization occurs was uncorrelated with climatic variables, but strongly correlated with initial litter chemistry (i.e. initial C and N concentrations). Based on comparisons of mineralization rates in regions receiving some of the highest atmospheric N inputs in the world and regions with very low deposition, Perez et al. (1998) suggested that the differences in mineralization rates were a consequence of the enhanced regional rates of N deposition, probably via longterm decreases in C/N ratios. For Sweden, FalkengrenGrerup et al. (1998) found higher potential net Nmineralization rates in areas of southern Sweden exposed to a total deposition of 17 kg N ha 1 yr 1 compared with areas exposed to 10 kg N ha 1 yr 1. Despite the relatively small differences in deposition, the N mineralization was 19.8 μmol N g 1 C week 1 in the region with the highest deposition compared with 9.5 μmol N g 1 C week 1 in the region with the lowest deposition. As the data was based on 600 sites with comparable climate and vegetation, Månsson & FalkengrenGrerup (2003) suggested that Ndeposition accumulation in the soil was the most likely cause for the reported differences in mineralization rates. Experiments with fertilizers support the findings of the deposition studies. Högberg et al. (2007) found strong negative correlations between gross Nmineralization and C/N ratio when investigating 26

29 a number of Swedish coniferous forest sites of differing productivity and exposed to N fertilisation. High N loading generally reduced the contribution by fungi to the microbial biomass (Högberg et al., 2007), since bacteria can maintain a relatively constant C/N ratio around 5 while fungi usually operate at higher C/N ratios (Paul & Clark, 1996). Based on their results, Högberg et al. (2007) suggested that the strong N sink that mycorrhizal fungi may constitute under low N conditions might change quickly when N increases, thus lowering the Nimmobilization capacity. According to the authors, this suggests a key role for fungi in regulating N cycling in boreal forests. Similar results with regard to N availability, mineralization rates, and the prevalence of fungi as compared with bacteria was found by Schröter et al. (2003). Several other longterm fertilization experiments besides Högberg et al. (2007) have also shown increased Nmineralization rates as a consequence of increased N availability (Magill et al., 1997; Lee & Caporn, 1998; Gundersen et al., 1998a). However, inconsistencies in the relationship between the soil C/N ratio and mineralisation and immobilization rates have been found. In the NITREXstudy, for example, the NH4 immobilization was highest at sites with low C/N ratios (Tietema, 1998). In shortterm experiments there is usually no increase in litter decomposition rates or Nmineralization rates with increasing N additions (Neuvonen & Suomela, 1990; Prescott, 1995). It is possible that prolonged N additions select for a microbial community adapted to more acid and Nrich conditions (Pennanen et al., 1998; Blagodatskaya & Anderson, 1999), a capacity that Månsson & FalkengenGrerup (2003) suggested have not had time to develop in shorter experiments. Other studies have suggested that the increase in gross N mineralization can only be partially explained by C/N ratios, and that it is instead due to a combination of an increase in the amount of organic matter, the quality of the organic matter, and the activity, biomass and composition of the soil microbial community (Zak et al., 2003; Manzoni et al., 2008). In accordance with the studies mentioned above, Eliasson & Ågren (2011), investigating the feedback from increasing soil inorganic N levels on N mineralization and tree growth in six boreal Scots pine stands in Sweden using the Qmodel, found that the mineralization of inorganic N decreased and N immobilization increased when more inorganic N became available due to slowly increasing N inputs over a century. However, when N was added as fertiliser, the amount of extra N was so large that it overrode the effect of decreased mineralization Effects of WTH WTH may influence decomposition and N mineralization not only by reducing the substrate availability but also by changing the abiotic environment. In general, soil temperature is increased in harvested plots as a consequence of a greater incidence of solar radiation resulting from the removal of the tree cover. The subsequent removal of logging residues might increase this effect even 27

30 further (PérezBatallón et al., 2001; Åström et al., 2005). Besides, soil moisture levels are generally lower where logging residues have been removed (PérezBatallón et al., 2001). Johnson & Curtis (2001) summarized the effects of CH versus WTH for 73 observations (from 26 publications) of temperate forest sites around the world. Their metaanalysis showed that forest harvesting had little or no effect on soil N across the entire data set. However, there were some significant effects. Residue removal caused a 6% reduction in Ahorizon N, whereas leaving residues on site caused an 18% increase compared to controls. According to Johnson & Curtis (2001), the positive effect on soil C and N of leaving residues on site seems to be restricted to coniferous species, since several studies have shown that residues had little or no effect on soil C and N in hardwood or mixed forests (Johnson & Curtis, 2001 and references therein). In some cases, the change was shortlived (less than 4 years) while in others it lasted longer (more than 18 years). In accordance with Johnson & Curtis (2001), Hopmans & Elms (2009) found large exports of N as a consequence of WTH for plantations of Pinus radiata in southern Australia, while Huntington & Ryan (1990), investigating the effects of WTH on soil N in a northern hardwood forest in the US, found no significant effect of WTH on the soil pool of N (although there were significant decreases in the concentrations in parts of the organic layer). In contrast to the above studies (where N decreased or was not changed as a consequence of WTH), Vanguelova et al. (2010), investigating a 28 year old second rotation stand of Sitka spruce growing in a peaty gley soil in the UK, found that WTH resulted in increased concentrations and stocks of total N compared with CH. The depletion of N in CH plots was attributed to higher soil mineralization rates where brash had been left. With regard to coniferous forests in Scandinavia, Saarsalmi et al. (2010) found a significant effect of WTH on the total amount of N in the organic layer in one fertile Scots pine stands in eastern Finland 22 years after harvest. However, in another less fertile Scots pine stand, there was no difference between WTH plots and control plots (Saarsalmi et al., 2010). That WTH had no effect on the N pools has also been noted in several other studies. Olsson et al. (1996) and Brandtberg & Olsson (2012) found no general effect of removing logging residues on soil N pools when investigating four coniferous sites in southern and northern Sweden 1516 and approximately 2526 years after harvest. Smolander et al. (2008), investigating a Norway spruce stand in central Finland ten years after harvesting, found that the mass loss of litter and the C/N ratio was not significantly affected by WTH. Wall (2008), also investigating a Norway spruce stand growing on moderately fertile soil in central Finland, found that removal of logging residues had no significant effect on the pools of organic matter and N in the Ohorizon or upper 10 cm of mineral soil four growing seasons after harvest. There are also studies where WTH have been applied but needles were left on site (Wall, 2008; Wall & Hytönen, 2011). As expected, these studies showed no difference between WTH and CH with regard to the pools of N in the upper part of the soil, since needles are the tree compartment that is richest in nutrients and leaving them on site provide a significant source of nutrients for trees. Furthermore, Rosenberg & Jacobson (2004), found no effect on soil N content when investigating the effects of slash removal after the second removal of logging residues in four wholetree thinned stands distributed across Sweden (one fertile Norway spruce stand in the southwest and three Scots pine stands in south, southcentral and central Sweden). The conclusion of Wall (2008), i.e. that in the short term, removal of logging residues do not impair pools of N in the soil nor site productivity on sites where the availability of N limits production, thus seems to be valid for most Scandinavian forest ecosystems. However, in contrast to the empirical Scandinavian studies mentioned above (where only soil N was taken into consideration), nutrient budget calculations (measuring net losses from the system, i.e. losses from both soil and trees) have shown that WTH will result in net losses of N from the forest ecosystems in large parts of Sweden (Akselsson et al., 2007). According to Akselsson et al. (2007), this implies that WTH can reduce the risk of eutrophication in the south where the N load is high (N accumulation rates of up to 14 kg ha 1 yr 1 with the deposition levels of 1998 and CH), but that it can also lead to N shortages in the north where the N load is considerably lower. A reduced N accumulation in southern and central Sweden as a consequence of WTH were also found by Hellsten et al. (2010) and Belyazid et al. (2008), using the same approach as in Akselsson et al. (2007). Besides, Belyazid et al. (2008) found similar results when using the dynamic model ForSAFEVEG to model the potential effects of WTH on concentration of organic N in humus until year The difference in concentration of organic N in humus between CH and WTH using ForSAFEVEG persisted under a changing climate (according to the IPCC scenario A2; Houghton et al., 2001). With regard to N mineralization, studies have shown that removal of logging residues either decreased the net Nmineralization in the long term (Piatek and Allen, 1999), had no effect on it (Vitousek et al., 1992; Brais et al., 2002), or resulted in an increase (PérezBatallón et al., 2001). In a study by Smolander et al. (2008), investigating a Norway spruce stand in central Finland ten years after harvesting, the mass loss of litter and the C/N ratio was not significantly affected by WTH. However, the mineralization of N and the amounts of C and N in the microbial biomass tended to be lower in WTH plots compared with CH plots, although not statistically significant. Furthermore, Smolander et al. (2008) found that the concentrations of total watersoluble phenols and an important group of phenols, condensed tannins, were both lower in the humus layer of WTH than in that of CH. The decay rates of litters from various plant species and the mineralization of C and N in the litter and humus layers have all been observed to correlate with the concentration of total soluble phenols. However, with the limited amount of information available, Smolander et al. (2008) emphasized that it cannot be concluded whether the decrease of soluble phenolic compounds due to WTH has any ecological consequences. What can be concluded is, according to the authors, that the decrease points to a change in the composition of organic matter as a consequence of WTH. A modelling study by Laurén et al. (2008) in a forested catchment in eastern Finland indicated that the removal of logging residues resulted in a reduced microbial immobilization of N and consequently, in no difference with regard to stream N export between CH and harvest with varying intensity of slash and SH. Litterbag experiments have demonstrated the immobilization of N into 28

31 forest debris at the early stages of decomposition (Berg & Söderström, 1979). According to Palviainen et al. (2004; 2010), logging residues in Norway spruce clearcuts can immobilize substantial amounts of N. Forest litter and logging residues may thus act as sinks of N (Vitousek & Matson, 1984), as soil bacteria, fungal mycelia and fauna take up soil N from outside the decomposing tissue Effects of SH In mature stands, stumps were found to account for around 815% of the N found in tree biomass (Palviainen et al., 2010 and references therein). Removal of stumps may thus influence decomposition, N mineralization and subsequent nitrification not only by reducing the substrate availability, but also by changing the chemical and physical environment. With regard to disturbance level, SH is often compared with mechanical site preparation (see for example Egnell et al., 2007). The effects of site preparation on soil nutrients are generally regarded to be rather shortlived (Walmsley & Godbold, 2010 and references therein) and not to result in longterm declines in site productivity (Örlander et al., 1996). However, in contrast to the belief mentioned above, the few studies of SH that do exist, have reported significantly reduced soil N pools as a consequence of SH. Zabowski et al. (2008), investigating five Douglas fir stands (Pseudotsuga menziesii (Mirb.) Fanco) stands across a range of conditions in the US, found that SH led to a 20% decline in mineral soil N over 2229 years as compared with areas where stumps had been left. Hope (2007) found that SH alone and in combination with soil scarification led to significant decreases in soil stocks of total and mineralizable N in the forest floor layer as compared with controls. Changes were apparent after one (mineralizable N and total N) and ten (total N) years. According to Hope et al. (2007), the lower soil N (and other nutrients) stock is a consequence of elevated rates of decomposition of surface organic matter following SH, rather than direct removal of organic material, as indicated by the substantial decrease in forest floor depth with time since harvest. Mineral soil N was not affected. Furthermore, coarse woody debris (CWD; of which stumps are the largest component normally left to decompose in Fennoscandian forests after clearcutting) have high C/N ratios (Palviainen et al., 2010). Various studies have thus suggested that N accumulation in CWD is an important mechanism of nutrient retention that might diminish nutrient leaching after disturbance and contribute to maintaining site fertility and productivity in Fennoscandian forests (Palviainen et al., 2010 and references therein). When investigating the C and N dynamics of Finnish Scots pine (Pinus sylvestris), Norway spruce (Picea abies) and Silver birch (Betula pendula) stumps that had been left to decompose for between 0 and 40 years, Palviainen et al. (2010) showed that the amount of N in stumps increased, indicating that external N accumulated in the stumps. After 40 years of decomposition, the amount of N was 1.7 and 2.7 times higher than the initial amount in pine and spruce stumps, respectively. N was released from birch stumps, but only after they had decomposed for 20 years or more. The results indicated that N is released considerably more slowly from stumps than from stems and branches of trees, and that stumps serve as N sinks (Palviainen et al., 2010). According to Palviainen et al. (2010), removing them may influence both N immobilization in the forest ecosystem and potential leaching of N into ground water and water courses after harvesting (see also section 6.2.6). Similar results were found in a study on stumps of Pinus radiata in New Zealand (Garrett et al., 2008). 5.2 Nitrification Nitrification converts NH 4 to NO 2 and NO 2 to NO 3. It has typically been associated with chemoautotrophic bacteria, although it is now realized that heterotrophic nitrification occurs and can be of significance, especially in forested soil (Paul & Clark, 1996). The chemoautotrophic nitrifiers are generally aerobs that derive their C largely from CO 2 or carbonates (Paul & Clark, 1996). Heterotrophic nitrifiers, on the other hand, are known to be capable of producing NO 3 from both inorganic and organic sources. In particular, heterotrophic fungi seem to have an organic pathway, converting amines and amides to NO 3 (Paul & Clark, 1996). Although N mineralization and nitrification are higher per gram of dry weight in the organic soil than in the mineral soil, the absolute levels expressed in grams of N in the soil are higher in the mineral soil (Lovett et al., 2004) Effects of temperature The temperature response of nitrification has an optimum in the range of 2035 C, but nitrification can occur even under snow cover, although very slowly (Paul & Clark, 1996). Temperature increases are known to result in increased potential nitrification rates in the rhizosphere of legumes (Singh et al., 2010) and according to Rennenberg et al. (2009), global warming might also be expected to promote nitrification across a wide range of forest ecosystems. However, results from field experiments are somewhat contradictory. Peterjohn et al. (1994) found no effect of increased temperature on net nitrification in a soil warming experiment in a hardwood forest in the US six months after the soil heating experiment had started. Shaw & Harte (2001) also reported no significant effects of temperature on gross or net nitrification rates when investigating a subalpine ecotone. However, Verburg et al. (1999) did find increased rates of net nitrification in the CLIMEX soil heating experiment in southern Norway. According to Rennenberg et al. (2009), it is unclear whether the different results reflect variable responses to warming due to different microbial communities being involved, or if other environmental controls override the effects of temperature. Generally, temperature, moisture and aeration (and sometimes also other environmental factors) interact, making up the seasonal effect of nitrification. In temperate areas, nitrification is thus often greatest in spring and fall and slowest in summer and winter (Paul & Clark, 1996). Another significant effect of temperature on nitrification may be through future increases in the frequency of freezingthawing 29

32 cycles. According to the review on N turnover in soil by Ollivier et al. (2011), increased nitrification rates have been observed after thawing in soils from moderate climatic zones Effects of moisture A factor of importance for nitrification is aeration, and the water status of the soil thus influences NO 3 production (Paul & Clark, 1996; Booth et al., 2005). 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 (Paul & Clark, 1996; Booth et al., 2005; Rennenberg et al., 2009). According to Booth et al. (2005), nitrification may respond positively to increasing soil moisture up to 0.01 Mpa, and then decline as the soil becomes saturated. Paul & Clark (1996), on the other hand, report that nitrification generally proceeds readily at the broader interval of 0.1 to 1 Mpa moisture tension and Dannenmann et al. (2006) found the optimum soil moisture to be in the range of 6070% of the waterholding capacity in a mountaineous beech forest in southern Germany. According to Rennenberg et al. (2009), climate change is expected to reduce soil water availability far beyond the optimum soil moisture content for nitrification in many parts of the world. Assimilation of NH 4 and NO 3 is generally more sensitive to both water stress and low temperatures than mineralization and nitrification. Inorganic N can thus accumulate in waterstressed or cool soils (Paul & Clark, 1996; Booth et al., 2005). However, NO 3 generally does not accumulate in mature forests unless there are no significant atmospheric sources of N (Paul & Clark, 1996) Effects of CO 2 As for the effect of temperature, the results with regard to the effects of elevated CO 2 on nitrification are also inconclusive. Nitrification has been reported to increase (Carnol et al., 2002; Barnard et al., 2005), decrease (Lagomarsino et al., 2008) or not change (Austin et al., 2009) in response to elevated CO 2. In a recent study, Schleppi et al. (2012) found that CO 2 enrichment accelerated net nitrification and induced higher NO 3 concentrations in soil water in a mature temperate forest in Switzerland that had been growing for eight years at elevated CO 2. Schleppi et al. (2012) suggested several explanations for the increase: 1) higher soil moisture as a consequence of reduced transpiration of CO 2 enriched trees, 2) enhanced N mineralization as a consequence of root exudates (priming effect) and/or 3) slightly decreased uptake of inorganic N as a result of a reduced N demand for fineroot growth Effects of N The competition for NH 4 ensures that nitrification depends primarily on the rate of ammonification, which in turn is strongly influenced by the C/N ratio of the substrate (Adams & Attiwill, 1993; Booth et al., 2005). N availability is thus a major controlling factor of nitrification, and nitrification is expected to increase in response to increased inputs of N into ecosystems. Accordingly, Barnard et al. (2005) reported gross and net nitrification in grasslands to increase as a consequence of N additions and Aber et al. (2003) found that net nitrification increased sharply below a treshold C/N ratio of between 20 and 25 in forest ecosystems. Emmett et al. (1998a) found a treshold of approximately 24 for a range of European sites. Studies by FalkengrenGrerup et al. (1998) and Booth et al. (2005) provide further evidence for the relation between C/N ratios and nitrification. However, there are also studies where no significant changes in nitrification rates as a consequence of increased N availability have been observed (for example Gundersen et al., 1998a). One possible explanation, suggested by Emmett et al. (1995), is that the link exists only if forests have experienced elevated N inputs over a prolonged period of time. That the response depends on changes in N pools rather than present day inputs was also suggested by Gundersen et al. (1998a), and several studies have demonstrated a delay or a transitory increase in nitrification of incoming NH 4 in coniferous and deciduous ecosystems without previous exposure to N (Fernandez & Rustad, 1990; Aber et al., 1993; Christ et al., 1995). One factor delaying the response of Ntransformation rates to N addition in low N ecosystems may be the slow change in the quality of soil organic matter and the associated soil biotic community (Emmett, 1999; Booth et al., 2005). Emmett et al. (1998b) suggested that Nrich soils have large active nitrifier populations that respond quickly to N inputs. LowN soils, on the other hand, have a low sensitivity to NH 4, and thus experience a lag period before stimulation of nitrification rates. Studies by FalkengrenGrerup et al. (1998) and Persson et al. (2000) support the necessity of an acclimation phase for the nitrifier community. They found that at sites subjected to high Ndeposition levels, the nitrification occurred at lower ph values than at sites with low N deposition. From these results, Persson et al. (2000) concluded that adaptations of nitrifiers to low ph can develop in soils with high NH 4 availability, but that the establishment of such a community takes time (at least 20 years) Effects of WTH There are few studies investigating the impact of WTH on nitrification. Vitosuek et al. (1992) found no effect of WTH as compared with CH on nitrification in a young loblolly pine plantation during five years following harvest. PérezBatallón et al. (2001), investigating the effect of WTH on a Pinus radiata plantation in Spain, on the other hand, found that nitrification was substantially higher in plots where slash had been removed as compared to plots where it had been left on site. However, the rapid development of grasses and shrubs resulted in low concentrations of NH 4 and NO 3 in the soil solution, leading PérezBatallón et al. (2001) to conclude that although nitrification rates might increase as a consequence of WTH, it does not necessarily lead to losses of N via leaching Effects of SH We are not aware of any scientifically published articles regarding the effects of SH on nitrification. 30

33 31

34 6. NITROGEN LOSSES 6.1 Denitrification and volatilization The release of N species to the air is the result of various transformation processes in the soil. Of the N species present in the soil, the noncharged forms of N, namely NH 3, NO, N 2 O and N 2, are able to volatilize. In the past, it was believed that denitrification was the only process causing loss of fixed N and gaseous N oxides. However, it is now known that these oxides are produced not only during denitrification, but also during N mineralization, N 2 fixation and nitrification (Paul & Clark, 1996). However, denitrification is still regarded as the main pathway for volatilization losses of N species. Denitrification can be both biotic and abiotic (socalled chemodenitrification). In forest soils, abiotic denitrification is far less important than biotic denitrification (Kimmins, 1997). Due to methodological difficulties in the quantification of denitrification, especially with regard to the dominant endproduct N 2, little is known about the actual rates of denitrification in forest soils (Rennenberg et al., 2009). According to Galloway et al. (2004), N 2 losses represent the largest uncertainty of the N cycle in forest ecosystems. Apart from N losses in the form of N oxides and N 2, N may also be lost to the atmosphere in the form of NH 3. If NH 4 is present in the soil in an unabsorbed state, such as in senescing or decaying vegetation or after application of fertilizer, NH 3 volatilization may occur (Paul & Clark, 1996). NH 3 volatilization is not described below, since it is unlikely to be of considerable importance in boreal forest soils Effects of temperature The temperature optimum of denitrification is thought to be similar to that of nitrification (Barnard et al., 2005). However, the range of temperatures under which denitrification may occur is considerably larger. The minimum temperature for denitrification is about 5 C, whereas the maximum is close to 75 C (Paul & Clark, 1996). Despite the strong theoretical temperature dependence of denitrification, not all studies have shown an effect of increased temperature on gaseous N losses. Melillo et al. (2002), for example, found no effects of ten years of increased temperature on gaseous N losses in a soil warming experiment in a midlatitude hardwood forest in the US, despite increased mineralization rates. However, Dannenman et al. (2008) pointed out that under field conditions, temperature effects on denitrification rates may often be masked by other environmental controls, such as ph, NO 3 concentration and soil water content. Furthermore, apart from its direct influence on microbial activity, Paul & Clark (1996) emphasized that temperature also affects the O 2 solubility and O 2 diffusion in water and, consequently, the denitrification potential. In accordance with the reasoning by Paul & Clark (1996), Öquist et al. (2004) showed that low temperature in a boreal forest soil promoted denitrification and N 2 O emissions by enhancing soil oxygen consumption and inducing anoxic conditions. In moderate climatic zones, one of the major future influences of temperature on denitrification might be through an increase in the number of freezingthawing events (something that may be expected in the future; Ollivier et al., 2011). Soil thawing is known to be accompanied by an accelerated release of nutrients and also by the emission of N 2 O and NO (Ollivier et al., 2011). Matzner & Borken (2008), reviewing the impact of freezethaw events on C and N losses from soils, found that N 2 O emissions from soils often increased after thawing. However, the increase lasted for a relatively short time (days to 12 months), and the greatest reported cumulative N 2 O emissions were around 2 kg N 2 ON ha Effects of moisture The enzymes involved in the microbial denitrification process are inhibited by oxygen. Consequently, O 2 availability in soil, and thus soil moisture, influence the denitrification process. The onset of denitrification occurs at about 60% waterfilled pore space and according to Paul & Clark (1996), roughly half of the variation in denitrification rates in the field can be explained by soil moisture. Although aerobic pathways do exist and denitrification may persist in rather dry soils when oxygen is depleted in soil microsites following high rates of soil respiration (Öquist et al., 2004), Rennenberg et al. (2009) concluded that the basic mechanistic understanding of denitrification implies that N losses from forest ecosystems will slow under drought conditions Effects of CO 2 Since most of the denitrification is accomplished by heterotrophic bacteria, the process is strongly dependent on C availability. According to Paul & Clark (1996), there is a general correlation between total SOM content and denitrification potential, and this improves if one considers only the supply of easily decomposable organic matter. Attempts to predict soil denitrifying capacity have shown that the amount of watersoluble C measurable in soil account for 71% of its denitrification potential (Paul & Clark, 1996). The increase in allocation of C to belowground as a consequence of elevated CO 2 (see Appendix A, section 4.3.3) thus implies a stimulation of denitrification at elevated CO 2. Apart from the extra C, Ollivier et al. (2011) suggested that the observed stimulation of denitrifiers under enhanced CO 2 concentrations might also be due to a general reduction of the redox potential in soil as a result of increased microbial activity. 32

35 There are indeed studies reporting increased N 2 O emissions in ecosystems exposed to increased levels of CO 2 (Singh et al., 2010 and references therein). However, there are also studies showing a decrease, as well as no effect at all. Since increases in N 2 O emissions were only observed when excess mineral N was available, Singh et al. (2010) suggested that when reactive N is high, the ecosystem will show increased N 2 O emissions under elevated CO 2. When reactive N is low, on the other hand, the reverse will happen. According to Singh et al. (2010), there is currently not much information available on such changes in N 2 O flux, but shifts in the structure of NH 3 oxidizing bacterial communities and decreases in their abundance under increased CO 2 levels have been observed Effects of N Fertilization with N may result in increased fluxes of gaseous N from the soil (Goulding, 1998; Mosier, 1998). It is wellknown that the use of N fertilizer in agriculture is a major source of anthropogenic N 2 O emissions (Singh et al., 2010) and Kimmins (1997) reported that up to 50% of the N that has been added to the soil in the form of fertilizer may be lost in a period as short as 24 days, if conditions are appropriate. Both mineralization and nitrification rates have been shown to increase as the C/N ratio of SOM is reduced (Gundersen et al., 1998; Aber et al., 2003). Mineralization and nitrification are known to be closely linked to N 2 O emissions (Firestone & Davidson, 1989) and increasing N concentrations can thus be expected to result in increased N 2 O emissions (Singh et al., 2010). In accordance with the theory, Klemedtsson et al. (2005) reported a strong exponential relationship between soil C/N ratios and mean annual N 2 O emissions from drained organic forest soils in Sweden. In these soils, the formation of N 2 O was found to be particularly pronounced in soils with C/N ratios lower than 25 (Ernfors et al., 2008) and in soils with high and fluctuating water tables (Bergkvist, 2007). That N fertilization may increase rates of N 2 O efflux also in nonorganic forest soils was shown by Klemedtsson et al. (1997), when investigating a N fertilized spruce stand in southwestern Sweden. However, the increase was only significant in the wetter part of the experimental area and emissions were generally low. Other studies in European forest ecosystems have also indicated that N fertilization might result in increased emissions of N 2 O from forest soils (Brumme & Beese, 1992; Mohn et al., 2000). A metaanalysis by Barnard et al. (2005) showed that N addition significantly increased field and laboratory emissions of N 2 O, although the effect of N addition on N 2 O efflux was not correlated to the amount of N added. In contrast to the studies referred to above, Wallenstein et al. (2006), investigating denitrification rates following five years of N amendment to a WTH forest plot included in the longterm soil productivity experiment in the US, found that fertilization did not affect insitu denitrification rates, despite evidence of accumulating NO 3 pools and higher activity of the denitrification enzyme in the fertilized plots. However, a NO 3 addition experiment in the laboratory resulted in increased N 2 O fluxes in soils from all plots, with the effect being strongest in the previously unfertilized plots. According to the authors, this suggested that chronic N fertilization has partially alleviated a NO 3 limitation of the denitrification rates. The authors also concluded that denitrification is thermodynamically constrained in these soils during summer, though there remains a potential for high activity under favourable conditions Effects of WTH We are not aware of any scientifically published articles regarding denitrification where the effects of WTH are compared with CH. In Wallenstein et al. (2006), WTH is compared with unharvested control plots and WTH where fertilizer has been applied. In this study, WTH resulted in a substantial reduction in denitrification rates compared with the unharvested control. On WTH plots where fertilizer was applied, on the other hand, denitrification rates remained high Effects of SH We are not aware of any scientifically published articles regarding the effects of WTH on denitrification. 6.2 Leaching Northern forest ecosystems have generally been considered to be N limited (Tamm, 1991), a state which is characterized by low leaching of inorganic N from growing forests. Leaching might, however, be induced by disturbances such as clearcutting (Adamson & Hornung, 1990; Wiklander et al., 1991; Rosén et al., 1996; Futter et al., 2010) or if the N input exceeds the capacity of the ecosystem to retain N (Aber et al., 1989; Dise & Wright, 1995; Dise et al., 1998; Moldan et al., 2006). Organic N, NH 4, and NO 3 may all leach from forest ecosystems. Often, dissolved organic N (DON) dominates N leaching from forest ecosystems (Smolander et al., 2001; Brookshire et al., 2007). In a Norway spruce stand in southeastern Finland, for example, DON comprised 6283% of total N in the soil solution over a period of five years (Smolander et al., 2001). Several studies have shown that in unpolluted temperate forests, DON dominate over inorganic N, while the reverse is the case in forests subjected to chronic N pollution (Brookshire et al., 2007 and references therein). The amount of N that is leached may vary substantially, from close to zero up to 50 kg N ha 1 yr 1 in areas with high external N loads (Dise et al., 1998). In Sweden, concentrations of NO 3 N in soil water at 50 cm depth have been shown to vary between <0.1 and >2 mg l 1 (Figure 3; PihlKarlsson et al., 2011). There is still relatively little information available with regard to how various environmental factors affect N leaching Effects of temperature Melillo et al. (2002) found no effect of ten years of increased temperature on leaching of organic and inorganic N in a soil warming experiment in a midlatitude hardwood forest in the US, despite increased mineralization rates. An increase in N mineralization 33

36 (which is commonly observed as a consequence of increased temperatures, see section 5.1.1) does not necessarily lead to additional N loss from the soil, as N uptake by roots or immobilization by microbes may occur. However, increased net mineralization increase the potential risk of N loss and there are examples of soil warming resulting in increased concentrations of NO 3 and NH 4 in runoff (Lükewill & Wright, 1997). Matzner & Borken (2008), reviewing the effect of freezethaw events on C and N losses from soils, found that seven out of eight studies on solute Nfluxes reported increased fluxes after exceptional soil frost events. In a field study of an alpine soil, NO 3 fluxes increased substantially after exposure to deep soil frost (11 C) in relation to sites of moderate frost (5 C). As a result, 11.4 kg N ha 1 were lost from the deep frost site during snowmelt, relative to 2.7 kg N ha 1 from the moderate frost site. Matzner & Borken (2008) hypothesized that their results were due to reduced NO 3 uptake by roots rather than increased net Nmineralization Effects of moisture According to Bechtold et al. (2003), two sets of circumstances could result in increased NO 3 outputs. First, leaching is limited by the availability of NO 3, which is released from organic matter. Alternatively, soil NO 3 could accumulate over longer time periods and leaching might then be limited by the availability of transport. Although hydrological factors are important in controlling N fluxes, studies on the effects of rainstorms and localized floodings on N leaching are relatively rare. One exception is the study by Bechtold et al. (2003) where a temperate floodplain alder ecosystem was investigated. They found that NO 3 was rapidly leached from soils during actual and simulated rainstorms. Localized flooding and direct leaching of streamside soils into surface waters were the main pathways of N leaching. While soil NO 3 pools were rapidly depleted during rainstorms, accumulation of soil NO 3 occurred over summer dry periods. That hydrological flow paths can override both abiotic and biotic retention mechanisms of N during the nongrowing season or during years with high rainfall, and that there can be direct flushing of N from the organic horizons into streams via horizontal flow was also shown by Dittman et al. (2007). Michalzik et al. (2001), reviewing 42 studies of forest ecosystems in the temperate zone, reported that the fluxes of DON in forest floor leachate increased with increasing annual precipitation Effects of CO 2 Studies of N leaching from forests subjected to elevated CO 2 are rare, but a recent study by Schleppi et al. (2012) showed that CO 2 enrichment decreased DON concentrations in soil solution in a mature temperate deciduous forest in Switzerland over eight years of treatment, compensating for a simultaneous increase in concentrations of NO 3. In the CLIMEXproject in Norway, where both temperature and CO 2 were manipulated to simulate the effects climate change on a boreal forest ecosystem consisting of a mixture of pine (Pinus sylvestris) and birch (Betula pubescens), soil N mineralization increased and the growing season was substantially prolonged in the climate change plot as compared with control plot, resulting in increased levels of inorganic N in runoff (despite lower N deposition due to exclusion of most dry deposition by the glasshouse). However, although the NO 3 export from the climate change plot was double that of the control plot, the fraction of the extra mineralised N that ended up in the runoff was small. The flux of organic N in stream water did not change significantly (Van Breemen et al., 1998). Figure 3. The concentration of nitraten (NO 3 ) in soil water at 50 cm depth at 60 forest sites included in the Swedish throughfall monitoring network. Modfied from PihlKarlsson et al. (2011) Effects of N Several fertilization studies have indicated that leaching may be induced by increased availability of N (Nilsson et al., 1988; Berdén 34

37 et al., 1997; Emmett et al., 1998a,b; Gundersen et al., 1998a). However, leaching only occurs if the N input exceeds the capacity of the ecosystem to retain N (Aber et al., 1989; Dise & Wright, 1995; Dise et al., 1998; Moldan et al., 2006), something that is generally true only for sites with very large N inputs (Emmett, 1999). Recent studies in several European countries have indicated that NO 3 leaching is one of the last things to occur in the succession of events related to increased N availability in an ecosystem. Although species change is apparent much earlier, the plant community still acts as a sink for N (Emmett, 2007). Ring et al. (2003; 2006) showed that N fertilization in two Swedish coniferous forest stands resulted in significantly increased concentrations of NH 4 and NO 3 in the soil solution, but without a corresponding increase in leaching. In one of the studies (Ring et al., 2003), where the forest had been clearcut before sampling (but fertilization occurred 7 years prior to harvest), concentrations of N in the fertilized plot decreased rapidly and was lower than in the control plot one year after harvest. The result can, according to the authors, be partly explained by the greater accumulation of N in the fieldlayer vegetation in the fertilized plot. In five field experiments distributed across Sweden, where N had been added at various frequencies, soil water samples from fertilised plots had higher average NO 3 concentrations (0.7 mg N l 1 ) than samples from the control plots (0.02 mg N l 1 ) (Bergh et al., 2008). However, the difference was not significant and values were generally low, as were the leaching rates ( kg N ha 1 yr 1 depending on frequency of fertilisation). When studying the effects of N deposition, Gundersen et al. (1998b) found that C/N ratios below 25 often coincided with elevated NO 3 leaching, whereas sites with C/N ratios above 30 seldom showed elevated leaching. Similar results were found by MacDonald et al. (2002), who demonstrated a significant relationship between N input by throughfall and NO 3 leaching, after dividing a European dataset into sites with C/N ratios below and above 25. After reviewing the impact of air pollution on Nleaching, Gundersen et al. (2006) later suggested that for coniferous forests, needle N content above 1.4% and (or) forest floor C/N ratio lower than 25 were tresholds for elevated NO 3 leaching. In contrast to Gundersen et al. (1998; 2006), Andersson et al. (2002) found no correlation between C/N ratio and N leaching when investigating seven Norway spruce sites distributed across Sweden. Instead, they demonstrated a direct correlation between soil flux density of N (N input net N mineralization) and NO 3 leaching. The C/N ratio as an indicator of NO 3 leaching has also been questioned by Moldan et al. (2006) and Emmett (2007), both of them referring to a N addition experiment in Gårdsjön, Sweden, that resulted in a significant increase in NO 3 leaching, without a corresponding change in the C/N ratio of the forest soil (Moldan et al., 2006). Emmett (2007) proposed competition for N between plants and microbes as a process being of importance for the amount of N leaching from forest soils. According to Brookshire et al. (2007), DON loss has previously been thought to be controlled by soil dynamics that operate independently of N supply and demand and thus track DOC. These results are supported by research showing that DON losses from water sheds remain consistent across broad geographic areas that vary considerably in N loading rates (Perakis & Hedin, 2002). Consistent with the results of the deposition study, Aandahl Raastad & Mulder (1999) and Sjöberg et al. (2003) reported that DON concentrations in soil solution at three Swedish spruce sites were not significantly affected by N fertilization, in short or long term. Furthermore, Michalzik et al. (2001), reviewing 42 studies of forest ecosystems in the temperate zone, found no relationship between the fluxes of DON in forest floor leachate and organic N stocks, soil C/N, litterfall or mineral N inputs. However, they did find a correlation between the fluxes of DOC and DON and that DON in forest floor leachates was positively related to DON in throughfall (variation in throughfall fluxes could explain 65% of the variation in DON fluxes from the forest floor). However, several more recent studies have shown large and persistent increases in soil solution DONflux as a consequence of longterm fertilization (Magill & Aber, 2000; McDowell et al., 2004; Fang et al., 2009). Pregitzer et al. (2004) found a more than sixfold elevation in the export of DON from four northern hardwood forests situated along a geographic gradient in the US. The leaching losses increased over time, which according to the authors suggested that N fertilization had altered the availability of the substrates that led to the formation of DON, or the processing of organic matter by soil microbial communities or both. When analysing soil and stream C and N in forest watersheds spanning a broad range of atmospheric N loading (545 kg N ha 1 yr 1 ) in the US, Brookshire et al. (2007) found distinct increases in DON loss with increasing N supply. The pattern was characterized by a steep initial increase followed by little change with additional N loading, suggesting that increases in DON losses are asymptotic and that once a treshold Ninput level has been exceeded, little changes in DON losses are to be expected. Recognizing that there are numerous other factors that may also influence soluble N pools and DON fluxes, Brookshire et al. (2007) concluded that the observed increases in DON losses in their study were driven by N supply. Furthermore, patterns in stream water N paralleled those in other studies (Hedin et al., 1995; Perakis & Hedin, 2002), showing a dominance of DON over inorganic N in unpolluted forests, and the reversal of this pattern as a consequence of chronic N pollution. That N losses may be strongly site and speciesspecific was emphasized in a study by Ladanai et al. (2007), when comparing a Scots pine and a Norway spruce stand in Central Sweden. While there were large losses of added N in the pine stand, more N than could be accounted for by inputs was recovered in the spruce stand Effects of WTH Several studies have suggested that removal of logging residues following clearcutting may reduce the accumulation of N in forest ecosystems subjected to high N load and thus potential N leaching (Lundborg, 1997), since removal of plant material reduces the pool of easily mineralizable N. According to Gundersen et al. (2006), WTH may increase the removal of N by 50 to 300% and lower leaching losses after clearcut as compared with CH. However, there are relatively few studies supporting these statements. In a 35

38 Sitka spruce forest in the UK, leaching of NO 3 was reduced by 90% after WTH as compared with CH (Emmett et al., 1991; Stevens et al., 1995). The effect was partly explained by a better establishment of grass when logging residues were removed by WTH. In accordance with the study by Emmett et al. (1991), Staaf & Olsson (1994) found that removal of slash in a clearfelled Norway spruce forest in southwestern Sweden led to reduced concentrations of NH 4 and NO 3 in soil water below the rooting zone. Within four years, though, treatment effects had disappeared. In contrast to the studies mentioned above, Wall et al. (2008) found no significant differences in the leaching of total and inorganic N compounds from the Ohorizon in WTH plots as compared to CH plots in a field experiment in Finland. Carlyle et al. (1998) even found that the leaching of N below the root zone decreased remarkably at some sandy Australian forest sites when woody residues of Pinus radiata (D.) Don. were added to the soil surface or buried into the topsoil. In coherence with the study by Carlyle et al. (1998), Mann et al. (1988) also observed higher NO 3 leaching after WTH than after CH. However, this was only apparent in the low fertility site investigated, while the opposite was true for the high fertility site. Using a spatially semidistributed FEMMA ecosystem model (tool for forestry environmental management), Laurén et al. (2008) found that removal of logging residues (slash removal of various intensities and slash removal in combination with SH) did not result in decreased N export to the stream in a forested catchment in eastern Finland. In the model, this result was explained by a decrease in microbial immobilization, which in the case of conventional clearcutting reduced net release of N and mitigated the export of N. Because the removal of logging residues decreased the decomposing N pool and simultaneously reduced microbial immobilization, no unambiguous hypothesis can, according to Laurén et al. (2008), be set in advance as to whether the removal of logging residues result in increased or decreased export of N to streams. Laurén et al. (2008) thus suggested that harvesting of logging residues in low atmospheric Ndeposition areas is not an effective strategy for reducing N loads in water bodies. Also Gundersen et al. (2006) suggested that it is not only the biomass harvest regime but also the sitespecific conditions that influence soil processes and subsequently controls N retention and leaching. Gundersen et al. (2006) also emphasized that WTH is often associated with other disturbances that may influence the magnitude of leaching losses, for example mechanical disturbance and compaction. Mechanical disturbance may delay regrowth of ground vegetation, something that was observed by for example Olsson & Staaf (1995). Furthermore, slash is often piled up on the site and N leaching can be substantial under slash piles (Rosén & LundmarkThelin, 1987; Staaf & Olsson, 1994). However, recent results from Wall (2008) indicated that the heaps of logging residues are minor sources of inorganic N Effects of SH Staaf & Olsson (1994), investigating the effects of SH in a clearfelled Norway spruce forest in southwestern Sweden, found that in the first year after felling NH 4 concentrations increased. High levels were noted for two years (but not higher than in the CH plots), followed by an additional twoyear period with elevated NO 3 concentrations. The latter was accompanied by higher levels of K, increased electrolytic conductivity and lower ph. According to Staaf & Olsson (1994), this suggested that decomposition and N mineralization of organic matter mixed into the mineral soil was stimulated, eventually resulting in increased rates of nitrification. The NO 3 concentration drop coincided with a rapid expansion of the ground vegetation, mainly consisting of Deschampsia flexuosa L. Trin., Carex pilulifera L. and Senecio species. After four years, all major differences between CH and SH had disappeared. 6.3 Tree harvest In a natural system, nutrients that have been taken up by the trees are returned to the soil when the trees die and their biomass decomposes. In a managed forest, however, various fractions (stems, slash and/or stumps) are removed from the system with the result that nutrients are permanently lost from the ecosystem. Akselsson & Westling (2005) have compiled data on N concentrations in stems, branches and needles of various tree species and used them for calculations of N losses at different harvest rates and scenarios. They found that the average concentration of N in stems were 1.1 mg g 1 for spruce and 0.9 mg g 1 for pine. The corresponding values in branches were 6.0 and 3.4 mg g 1 and in needles 11.2 and 12.4 mg g 1. Similar values were found by Alriksson & Eriksson (1998; 1.8 and 1.5 mg g 1 in stems, 5.1 and 4.3 mg g 1 in branches and 11 and 15 mg g 1 in needles of spruce and pine) and by Thelin (1998; and 1415 mg g 1 in needles of spruce and pine). From these data, it is obvious that the amount of N lost through harvesting depends not only on the amount of harvested biomass, but also on what fraction of the tree that is harvested and on tree species. The calculations by Akselsson & Westling (2005) showed that WTH resulted in a two times higher N loss than CH. Furthermore, the difference in N loss between CH and WTH is greater for spruce forests compared with pine forests, since spruce trees have a larger relative fraction of branches than pine trees (Akselsson & Westling, 2005). Mann et al. (1988) found that nutrient removals of N ranged from kg ha 1 when only stems were harvested in coniferous forest stands in the US. When WTH was applied, removals were in the range of kg ha 1. For hardwood stands the corresponding numbers were kg ha 1 and kg ha 1. In many cases, the removal of N doubled, or more, when WTH was applied (as compared with CH). When comparing direct removal of nutrients through harvest with hydrological N losses, removal of N in wood with WTH generally exceeded hydrologic nutrient losses (Mann et al., 1988). The effect of SH on N losses are not fully investigated, but preliminary results in Hellsten et al. (2010) show that even if the N loss is smaller than when slash is removed (N concentrations in stumps are lower than in slash), it is not negligible since the biomass of stumps is substantial. 36

39 7. CONCLUSIONS Recent advances have resulted in a more comprehensive understanding of how climate change and forest management practices affect processes controlling the N dynamics of boreal forest ecosystems. However, it is still difficult to make informed and wellbased predictions about the effects on an ecosystem scale level. The problem is not the amount of information available, but that available results are both straggling and inconclusive. In particular, the impact on processes controlling soil N and N losses are unclear. The information with regard to how these factors affect N input and plant responses are more easily interpreted, although there are uncertainties also here (for example with regard to uptake processes). One of the greatest challenges for the future is to attain longerterm, consecutive data sets on N cycling, where the effects of an environmental factor or management practice is evaluated at more than one point in time, and where the effects of various combinations of factors that act simultaneously on the ecosystem are investigated. Another challenge is to clarify the linkages between, and understand the coupling of, the N, C, P and base cation cycles. Without such information, predictions and conclusions with regard to the effects of climate change, N fertilization, WTH and SH on N cycling in boreal forest ecosystems are difficult. 37

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