IDENTIFICACIÓN DE LOS OBJETIVOS DE LAS AULAS VIRTUALES

different sources and then undergoes a series of biochemical reactions to continue its active cycling within the soil-plant-atmosphere loop. In contrast to C in plants, which has its main origin in atmospheric CO2, N in plants mainly originates from soil. The δ15N value of plants, however, is not exactly the same as that of total N (TN) in soil (Makarov, 2009). The difference reflects the range of bioavailable and non-bioavailable N compounds in bulk soil N, and also fractionation of 15N at the time of N uptake by plants. Within plants, there are also additional biological processes causing 15N discrimination and variability in plants (Evans, 2001; Makarov, 2009).

Figure 2-7: A simple model of N cycling in the soil-plant-atmosphere system.

2.1.3.1

Nitrogen Isotopic Composition of Source N

The δ15_{N value of soil N is one important factor affecting plant} δ15_{N (Vallano and Sparks,} 2012). The most common forms of N in soil taken up by plants are inorganic (nitrate (NO32-) and ammonium (NH4+)) and dissolved organic (e.g. simple proteins, amino acids and amino sugars) compounds (Emmerton et al., 2001; Kielland et al., 2006; Näsholm et al., 2009; Schimel and Chapin, 1996; Wei et al., 2013). There are differences in δ15_N among these sources because of different biochemical reactions in soil (N mineralization, nitrification, denitrification and volatilization) (Hogberg, 1997; Mariotti et al., 1981). As a result, bulk soil δ15_{N is not always a good representation of the} δ15_{N of source N to} plants because bulk soil might be dominated by non-bioavailable N (Hogberg, 1997).

The quantitative determination of the natural abundance of 15N across the range of soil sources is very difficult for at least a couple of reasons (Hogberg, 1997; Robinson, 2001).

First, turnover times for both soil inorganic (Davidson et al., 1992; Davidson et al., 1990; Hart et al., 1994) and organic forms (Kielland, 1995) are short, varying from a few days to weeks. Second, the complex dynamics of soil N makes the development of any protocol for accurate extraction and analysis of very small fractions of mobile N forms extremely difficult (Hogberg, 1997; Robinson, 2001). Generally, such studies are performed using 15N labeling to identify plant preference for different forms of N (Houlton et al., 2007; Schimidt and Stewart, 1997).

The isotopic fractionation (ε) associated with different N biochemical reactions in soils (Table 2-2) generally result in products more depleted of 15N than their substrates (Hogberg, 1997; Makarov, 2009; Robinson, 2001). Accordingly, the following order in δ15_{N for different N forms can be predicted:}

δ15_{Norg >} δ15_NNH4+ _> δ15_NNO3- _{(Makarov, 2009)}

While such a pattern has been reported for coniferous soils in Japan (Koba et al., 1998), deviations from this trend are common. The δ15_{N signal of the products of soil biological} transformations can be higher than their substrate because they can be a substrate for another biological reaction having a higher reaction rate. For example, δ15_NNO3- _{can be} higher than δ15_NNH4+ _{because of a larger degree of denitrification than nitrification} (Makarov, 2009).

Another factor adding to the complexity of the isotopic composition of bioavailable N in soils is the δ15_{N gradient with soil depth, particularly in natural ecosystems where soils} are not tilled (Hogberg, 1997). This can affect plants δ15_{N with different root} morphologies and rooting depth distribution (Makarov, 2009). A trend of increasing bulk soil δ15_{N with soil depth reported in many studies (Hogberg, 1997; Koba et al., 1998;} Makarov et al., 2008; Martinelli et al., 1999; Nadelhoffer and Fry, 1988) has been explained by the input of fresh litter depleted of 15N in topsoil, and the gradual accumulation of decomposed OM enriched in 15N with depth (Makarov et al., 2008).

Table 2-2: N isotopic fractionation for different biochemical reactions in the N cycle (from Robinson, 2001 and Houlton and Bai, 2009).

Process ε (‰)

Microbial N2 fixation (N2 Organic N) 0-6

Nitrification (NH4 + NO3 - ) 15-35 NH3 volatilization (NH4 + (aq) NH3(g)) 40-60 Denitrification (NH4+ N2/N2O/NO2) 28-33 Plant NO3 -

uptake and assimilation (NO3 -

Organic N) 0-19

Plant NH4 +

uptake and assimilation (NH4 +

Organic N) 9-18

Microbial N assimilation (NH4+ Organic N) 14-20

Ammonification (mineralization) (Organic N NH4 +

) 0-5

Nitrate Leaching 0-1

Nonetheless, no regular trend of δ15_{N for inorganic N with soil depth has been reported;} some studies report an increase in δ15_NNH4+ _and δ15_NNO3- _{with depth (Koba et al., 1998),} while others report a decrease in δ15_NNH4+ _{with depth over a wide range of alpine and} tundra soils (Makarov et al., 2008). Moreover, much evidence supports diversity in the ability of plants to acquire N depending on their rooting depth and phenology (Robinson and Rorison, 1983; Shaver and Billings, 1975), life form (trees, shrubs and herbs) (Schulze et al., 1994) and preference for different forms of N (Houlton et al., 2007; Miller and Bowman, 2002; Nadelhoffer et al., 1996) at different times of the year (Shaver and Kummerow, 1992). In short, a wide range in soil δ15_{N among N forms and in the type of} N acquired by plants should be considered when interpreting isotopic signals of plants from different ecosystems.

Mycorrhizal fungi association with plants is another factor affecting the δ15_{N of source N} for plants. Approximately 80 % of terrestrial plants and more than 90 % of vascular plants form biological symbioses with soil mycorrhizal fungi (Wang and Qiu, 2006). This association increases their ability to capture soil nutrients, such as nitrogen (Leigh et al., 2009) and phosphorous (Bolan, 1991). Plants are more reliant on mycorrhizal fungi for N acquisition under conditions of low N availability (Craine et al., 2009). Therefore, plant-

fungi symbiosis is of particular importance in Arctic and subarctic ecosystems where inorganic N is limited (Hobbie and Hobbie, 2008).

Arbuscular mycorrhizal fungi, which mainly colonize the roots of herbaceous plants and tropical trees, play an uncertain role in modifying the δ15_{N of plants. It has been} suggested that they lack the enzymatic capability to mineralize soil organic N (Hobbie and Hobbie, 2008), and are involved mainly in phosphorous plant nutrition (Craine et al., 2009). In contrast, ectomycorrhizal and ericoidmycorrhizal fungi, which colonizing mainly the roots of trees and shrubs, have the enzymatic capability to dissolve organic N in soils and provide it for plants (Abuzinadah and Read, 1986; Emmerton et al., 2001; Hobbie and Hobbie, 2008). A global survey of plant δ15_{N shows that arbuscular} mycorrhizal, ectomycorrhizal and ericoidmycorrhizal plants are depleted of 15N by 2 ‰, 3.2 ‰ and 5.9 ‰, respectively, relative to nonmycorrhizal plants (Craine et al., 2009; Michelsen et al., 1996, 1998). The cause of such isotopic differences between non- mycorrhizal plants and plants with different types of mycorrhizal associations is not well understood. Some studies explain this pattern by greater uptake of 15N-depleted organic N from soils by ericoidmycorrhizal plants and greater use of inorganic N by non- and ectomycorhizal plants (Michelsen et al., 1996). Other studies reject this hypothesis and suggest that the δ15_{N of soil inorganic N is insufficiently low to explain this pattern.} Instead, they attribute lower δ15_{N of mycorrhizal plants to discrimination against} 15_N during transfer of N from fungi to host plants (Hobbie and Hobbie, 2006; Schimidt and Stewart, 1997) or to preferential transfer of 15N-depleted N compounds (e. g. glutamine) produced during enzymatic reactions within fungi to the plant partner (Hobbie and Hobbie, 2008; Hobbie et al., 1999).

There are also a group of bacteria (rhizobia) that make an association with specific groups of plants (legumes) to fix atmospheric N2. This symbiosis affects host plants δ15N (consistently around 0 ‰) by providing them with N sources having δ15_{N close to those} of atmospheric N2 (δ15NAIR = 0 ‰) (Mariotti, 1983). There is negligible fractionation of 15

N during biological fixation of N by these bacteria (Kohl and Shearer, 1980). None of the plants considered in the present study, however, are from the legume family.

Herbivory activities (grazing, trampling, excretory products and soil disturbance) also can change δ15_{N of soil and plants. Herbivory can affect N dynamics in soils and plants by} changing N availability (Frank et al., 2000), altering the rate of various soil N processes (Frank et al., 2000; Wolf et al., 2010), and modifying litter quality and plant composition (Augustine and Frank, 2001; Semmartin et al., 2004). These effects have been suggested to be different among different ecosystems, functional groups and species (Augustine and Frank, 2001; Zheng et al., 2012), but no systematic pattern has emerged. Some studies have reported 15N enrichment of soils and plants with increased herbivory activities (Aranibar et al., 2008; Coetsee et al., 2010; Frank and Evans, 1997; Li et al., 2010), whereas others report depletion of 15N (Frank et al., 2000; Golluscio et al., 2009) or little or no effect (Cook, 2001; Wittmer et al., 2010; Xu et al., 2010).

2.1.3.2

Nitrogen Isotopic Fractionation during N Uptake and

Assimilation in Plants

Nitrogen isotope fractionation during plant uptake of NO3- and NH4+ is generally considered to be controlled by two factors: (i) external N concentration (Evans, 2001; Hogberg, 1997; Kolb and Evans, 2003), and (ii) efflux of 15N-enriched inorganic N and/or 15N-depleted organic N from roots after N uptake when reduction/assimilation occurs in the roots (Craine et al., 2015; Evans, 2001; Robinson, 2001). Many experiments have reported negligible ε of N during plant uptake under low [NO3-] and [NH4+] (~0.5 mol m3-), with an increase in ε under higher concentrations (Mariotti et al., 1982; Yoneyama et al., 2001). Because most terrestrial ecosystems are N-limited (Thomas et al., 2013; Lebauer and Treseder, 2008), discrimination against 15N during plant uptake is probably negligible under most natural conditions (Evans et al., 1996; Evans, 2001).

Once NO3- and NH4+ are taken up from soils by plants, they undergo various reduction processes to be assimilated. Assimilation of NH4+ by glutamine synthetase-glutamate synthase (GS-GOGAT) occurs in the roots close to the site of uptake; this arrangement limits toxic accumulation of NH4+, which interferes with energy metabolism and ATP production (Hopkins and Hüner, 2009). NO3- can be assimilated in both roots and leaves through a series of reduction reactions that first reduce NO3- to NO2- and then NO2- to

NH4+ through the nitrate reductase (NR)-nitrite reductase (NiR) pathways. The NH4+ then follows the GS-GOGAT pathway to be converted into amino acids and other organic forms of N in plants (Evans, 2001; Hopkins and Hüner, 2009). These enzymatic reactions are the major steps of 15N fractionation in plants. Isotopic fractionations of ‒15 ‰ and ‒17 ‰ have been reported for NR and GS, respectively (Evans, 2001).

2.1.3.3

Intra-plant Variation in δ

_N

There are many factors contributing to intra-plant variation of δ15_{N including: (i)} variation in plant nitrogen sources as different organs form and expand, (ii) different patterns of N assimilation with either NO3- or NH4+ as the primary N source, (iii) reallocation and transportation of N macromolecules between sink and source organs, and (iv) organ-specific efflux of N (Evans, 2001; Szpak et al., 2013).

In contrast to δ13_{C, which shows more negative values in leaf than stem or root tissues,} leaves normally have higher δ15_{N than other organs, particularly roots (Evans et al., 1996;} Evans, 2001; Yoneyama and Kaneko, 1989). Kolb et al. (2002), however, has reported the opposite pattern, with roots having higher δ15_{N than leaves and stems in two species} of deciduous trees. When NO3- is the sole N source, significant intra-plant variation occurs, with leaves having much higher δ15_{N than roots (up to 7 ‰) (Bergersen et al.,} 1988; Yoneyama and Kaneko, 1989). This likely reflects different patterns of NO3- vs. NH4+ assimilation. Whereas NH4+ is assimilated right after uptake in roots, some NO3- is assimilated in roots while the remaining unassimilated NO3-, which is now more enriched in 15N, is transported to shoots to be the precursor for N assimilation in leaves (Evans, 2001).

Reallocation of N through different enzymatic reactions and then transport of products to different organs also causes δ15_{N intra-plant variations. This occurs because all enzymatic} reactions involved should produce molecules with lower δ15_{N than the original source} (Yoneyama et al., 1998). Loss of NH3 through plants leaves and efflux of organic N from roots also can enrich these organs in 15N (Evans, 2001; Shearer and Kohl, 1986).

2.1.3.4

Environmental Factors and Plant δ

_N

Among different environmental factors, the significant roles of Mean Annual Temperature (MAT) and MAP in controlling soil and plants δ15_{N signals have been} reported in many studies (Amundson, 2003; Austin and Vitousek, 1998; Craine et al., 2009; Ma et al., 2012). A systematic local and global decrease in soil and plants δ15_N with increasing MAP and decreasing MAT has been observed by Amundson et al. (2003) and confirmed by Craine et al. (2009) for ecosystems with MAT ≥ ‒0.5 ᵒC. This relationship may be related to the change in N cycling in soil and plants, the rate of soil N transportation (e.g. denitrification, volatilization) and/or dependence on mycorrhizal fungi (Craine et al., 2009). Changes in the amount of rainfall and subsequently soil water availability can affect the openness of the N cycle (Austin and Vitousek, 1998; Schulze et al., 1991), with a more open N cycle in drier sites most probably reflecting a higher rate of gaseous loss of N (volatilization and denitrification) from soil and plant systems. These reactions leave the soil N pools more enriched in 15N due to the large ε associated with them (Table 2-2) (Hogberg, 1997).

One might expect low denitrification rates in drier sites due to the lack of anaerobic conditions required for this reaction. What mainly limits denitrification, however, is NO3- availability (Burchill et al., 2014; Groffman et al., 1993), which is not limited in drier sites characterized by more nitrification. Even temporary soil waterlogging after a heavy rainfall can provide favorable conditions for denitrification in drier sites. In addition, lower N availability in wetter sites (Schuur and Matson, 2001), more reliance of plants on mycorrhizal fungi for N acquisition (Craine et al., 2009), and/or a complete conversion of soil NO3- to gaseous forms of N through denitrification ‒ which does not leave the system enriched in 15N (Houlton et al., 2006) – might contribute to lower soil and plants δ15_{N in} wetter ecosystems. It has also been suggested that greater N availability in drier sites due to less plant N demand, can stimulate NH4+ volatilization and therefore higher δ15N of soil and plant (Austin and Vitousek, 1998). N loss reactions (denitrification and volatilization) should mainly affect actively cycling pools of N, given that identical soil

and plant N isotopic responses have been observed along precipitation gradients (Austin and Vitousek, 1998).

What is clear from these studies is that N cycling in ecosystems, and in turn soil and plant δ15_{N, are highly responsive and sensitive to climatic factors. Moreover, these changes in} δ15_{N of plants can be tracked along trophic levels from primary producers to consumers} (animals and humans) (Schwarcz, 1999; Szpak et al., 2010).