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A plant perspective on nitrogen cycling in the rhizosphere
D. Moreau1, R.D. Bardgett2, R.D. Finlay3, D.L. Jones4,5 and L. Philippot1*
1Agroécologie, AgroSup Dijon, INRA, Univ. Bourgogne, Univ. Bourgogne Franche-
Comté, F-21000 Dijon, France
2The Univ Manchester, Sch Earth & Environm Sci, Michael Smith Bldg, Manchester M13
9PT, UK
3Swedish Univ Agricultural Sciences, Uppsala Biocenter, Dept Forest Mycology and Plant
Pathology, Box 7026, SE-750 07 Uppsala, Sweden
4Bangor Univ, Environment Centre Wales, Deiniol Rd, Bangor LL57 2UW, Gwynedd, UK
5UWA School of Agriculture and Environment, University of Western Australia, Crawley,
WA 6009, Australia
Abstract
1) Nitrogen is the major nutrient limiting plant growth in terrestrial ecosystems, and the
transformation of inert nitrogen to forms that can be assimilated by plants is mediated by soil
microorganisms.
2) The last decade has witnessed many significant advances in our understanding of plant-
microbe interactions with evidence that plants have evolved multiple strategies to cope with
nitrogen limitation by shaping and recruiting nitrogen-cycling microbial communities.
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However, most studies have typically focused on the impact of plants on only one, or
relatively few, processes within the nitrogen cycle.
3) This review synthesizes recent advances in our understanding of the various routes by which
plants influence the availability of nitrogen via an array of interactions with different guilds
of nitrogen-cycling microorganisms. We also propose a plant-trait based framework for
linking plant N acquisition strategies to the activities of N-cycling microbial guilds. In doing
so, we provide a more comprehensive picture of the ecological relationships between plants
and nitrogen-cycling microorganisms in terrestrial ecosystems.
4) Finally, we identify previously overlooked processes within the nitrogen cycle that could be
targeted in future research and be of interest for plant health or for improving plant nitrogen
acquisition, while minimizing nitrogen inputs and losses in sustainable agricultural systems.
Introduction
Like all living organisms, plants need nitrogen (N), an essential component of nucleotides and
proteins. N also forms the skeleton of chlorophyll and is one of the major plant
macronutrients. Despite being abundant, most N in the atmosphere is in the inert form N2,
which is not directly useable by plants. As such, N is the most common limiting nutrient for
plant growth (LeBauer & Treseder 2008). The production of N fertilizers by the energy-
demanding Haber-Bosch conversion of atmospheric N2 to ammonia has increased by a factor
of 20 since 1950 (Glass 2003) and the amount of N2 artificially fixed to produce fertilizers
(about 9.5 × 1012 mol per year), now exceeds the total biological N fixation in terrestrial
systems (7.5 × 1012 mol per year) (Canfield, Glazer & Falkowski 2010). This excess reactive
N from fertilizer use poses a threat to soil, water, and air quality due to nitrate leaching and
emission of N2O, a potent greenhouse gas that is also involved in the destruction of the ozone
layer (Montzka, Dlugokencky & Butler 2011). There is thus a critical need to increase the
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efficiency of plant N acquisition and reduce reliance on N fertilizers in sustainable
agricultural systems (Philippot & Hallin 2011; Sutton et al. 2011; Jones et al. 2013).
How plants respond to spatial and temporal variation in nutrient availability has been a
topic of long standing interest in ecology. A key outcome of this interest has been recognition
of the existence of an evolutionary trade-off between fast-growing exploitative and slow-
growing conservative growth strategies (Grime 1977; Chapin & Stuart 1980; Wright et al.
2004). To better categorize plant distribution along resource gradients in relation to these
strategies, trait-based approaches are now commonly used (Cornwell & Ackerly 2009;
Ordoñez et al. 2009; Moor et al. 2017). However, the focus of these trait-based approaches
has recently shifted below-ground, from leaf to root traits and associations with soil microbial
communities and nutrient cycling processes (Laliberté 2017; Bardgett 2018). This reflects
knowledge that (i) soil-N availability in soil is strongly dependent on various microbial guilds
transforming inert N into N compounds that can be assimilated by plants. and (ii) plants can
shape the activity and the composition of soil microbial communities (Berendsen, Pieterse &
Bakker 2012; Philippot et al. 2013). Since several different microbial guilds are of importance
for N availability in soil, it therefore seems reasonable to hypothesize that plants have also
evolved multiple mechanisms to acquire N by shaping and recruiting these N-cycling soil
microbial communities.
While the influence of plants on N-cycling microorganisms has previously been
synthesized, to date these reviews have largely focused on a few processes within the N cycle
(Subbarao et al. 2009; Kuzyakov & Xu 2013; Coskun et al. 2017). We therefore lack a
comprehensive view of how plants influence the different microbial N transformations to
increase N availability and retention in the plant–soil system, which would support the
hypothesis of multiple plant strategies for controlling key steps in the microbial N cycle. The
overall goal of this paper is to provide such a holistic view of ecological relationships between
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plants and N-cycling microorganisms in terrestrial ecosystems, and propose a plant-trait based
framework for linking plant N acquisition strategies to the activities of N-cycling microbial
guilds. We draw upon a growing set of studies investigating the mechanisms and processes
involved, the consequences for microbial communities, and the ultimate fate of the N they
process. Finally, we identify processes that could be of interest in the future in order to
improve plant N acquisition and health, while minimizing N inputs and losses in sustainable
agriculture.
What nitrogen forms are taken up by plants?
To understand how plants influence microbial N transformations to improve their nutrition
requires knowledge about the different N forms that plants can take up, and how uptake of
these compounds is influenced by both their availability in soils and by plant properties
(Figure 1). Soluble N in soil varies greatly over space and time, and most plants appear to be
highly opportunistic in exploiting any ephemeral micro-scale patches of N (Chapin, Matson &
Mooney 2002). The inorganic forms of N absorbed by plants are generally nitrate and
ammonium (Courty et al. 2015). Uptake of nitrate is an active, energy-requiring process
because it must be absorbed against an electrochemical gradient (Haynes & Goh 1978). It is
therefore generally less energetically efficient than ammonium uptake (Wang et al. 1993),
which can be either active or passive, depending on the ammonium concentration in soils.
Nitrate assimilation is also energetically expensive because a preliminary reduction of nitrate
to ammonium is required, while absorbed ammonium can be directly assimilated in order to
synthesize amino acids and other organic compounds. On the other hand, since it does not
bind to the cation exchange complex of soils, nitrate is more mobile than ammonium in soils,
and therefore usually more available for plant N uptake (Courty et al. 2015).
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More recently, plants have also been shown to take up N in different organic forms (Nasholm
et al. 1998; Jones et al. 2005), that are the least energy-expensive for plants since the
assimilation process is bypassed (Chapin, Matson & Mooney 2002). Plants can take up and
thrive on a wide range of dissolved organic N forms, ranging from simple, low molecular
weight compounds (e.g. amino acids, oligopeptides, nucleotides, urea) through to more
complex polymeric moieties (e.g. proteins) (McKee 1962; Paungfoo-Lonhienne et al. 2008;
Hill et al. 2011). At high exogenous concentrations (> 1 mM), uptake of soluble organic N
into the roots can occur by passive diffusion. However, transport across the plasma membrane
is generally driven by families of H+-ATPase-fuelled active transporters, the genes for which
appear to be constitutively expressed at high levels, irrespective of soil N supply (Jones &
Darrah 1994). These high affinity transporters are capable of depleting external solute
concentrations to the nM level, while maintaining cytoplasmic concentrations at the mM
level, demonstrating the efficiency of these N capture systems (Jones & Darrah 1994). A
negative consequence of maintaining this steep concentration gradient across the plasma
membrane is the continual passive loss of organic N from the root back into the apoplast and
soil, via rhizodeposition (Phillips, Fox & Six 2006). This has led to speculation that the
uptake of organic N from outside the root is not associated with acquiring new soil-derived N,
but rather provides a mechanism to recapture N from the plant to prevent excessive microbial
proliferation in the rhizosphere (Jones, Nguyen & Finlay 2009). Alternatively, the expression
of these transporters may be related to the exchange of signaling molecules between plant
growth promoting microorganisms and roots (e.g. oligopeptides) or in the root sensing of their
environment (e.g. hotspots of nutrient turnover) (Contreras-Cornejo et al. 2016).
Despite overwhelming evidence that roots can take up organic N, the ecological
significance of this N acquisition pathway remains controversial. Soil solution contains
thousands of individual organic N compounds, however, most of these are thought to be
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comprised of humic-rich by-products of microbial breakdown and to be relatively unavailable
to plants (Warren 2014; Warren 2017). In addition, simple compounds which are not widely
used in plant metabolism but which may be common in soil may also not be taken up by roots
(e.g. glucosamine, a potential breakdown product of fungal cell wall chitin) (Roberts & Jones
2012). Many other compounds are present at very low concentrations (<1 µM), which might
imply that they are of limited significance for plant N nutrition as they are close to the net
influx-efflux equilibrium point. What is critical, however, is not their concentration in soil
solution, but their rate of replenishment and the potential reserves that exist in the soil
exchange phase. For example, it has been estimated that the soil solution amino acid pool is
maintained at low concentrations in all ecosystems (ca. 20 µM), but is replenished over 1000
times a day (Glanville et al. 2016). This rate of cycling can be orders of magnitude faster than
the rates of NH4+ and NO3
- production in soil (Jones & Kielland 2002; Andresen et al. 2015)
Although there have been significant advances in the understanding of the different
forms of N that can be taken up by plants, fewer studies have investigated how the uptake of
these N-forms is differentially modulated by abiotic and biotic factors. In the future, gaining
more detailed information on the relative importance of organic versus mineral N sources for
plants in agricultural systems, especially in organic systems (where no mineral N inputs are
used) is crucial in order to better exploit microbial processes in the context of reducing
chemical inputs. Moreover, better understanding of the effects of these numerous factors,
especially of their combinations, may help to improve understanding of plant-microbe
interactions, both in terms of competition and cooperation. Unlike almost all previous studies,
these plant N uptake experiments need to be designed in such a way that the outcome is not
biased by the experimental conditions (Hill & Jones 2019).
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Plants acquire nitrogen by recruiting microorganisms through mutualism and
multitrophic interactions
The vast majority of terrestrial plants have evolved different mutualistic associations
with fungi and bacteria, involving physical integration of symbiotic partners within
specialized root structures. The most ancient and widespread form of plant symbiosis
(involving 70-90% of extant species) involves arbuscular mycorrhizal (AM) fungi belonging
to the phylum Glomeromycota (Parniske 2008) and is thought to have originated 460 million
years ago. Up to 20% of the photosynthetically derived C can be allocated to fungal mycelia
forming up to 100 m of hyphae per cubic cm of soil and capable of accessing soil pores that
are inaccessible to plant roots. The view that AM fungi only contribute to phosphorus
acquisition for their host plant has recently changed (Hodge & Storer 2015), and there is now
evidence that AM fungi not only take up nitrate, ammonium and organic N, but also transfer
N to the plants (Johansen, Finlay & Olsson 1996; Hodge, Campbell & Fitter 2001; Toussaint,
St-Arnaud & Charest 2004; Govindarajulu et al. 2005; Tanaka & Yano 2005). It has also been
shown that this N uptake by the fungal symbiont and transport to the root is stimulated by C
supply from the host plant across the mycorrhizal interface (Fellbaum et al. 2012).
In contrast to the AM fungi, ectomycorrhizal and ericoid mycorrhizal fungi produce
degradative enzymes that can decompose more or less recalcitrant organic compounds
containing polymers of N (Read & Perez‐Moreno 2003). Uptake and assimilation of
ammonium by ectomycorrhizal mycelium and translocation of N in the form of amino acids
within the ectomycorrhizal mycelium have been demonstrated in different ectomycorrhizal
species (Finlay et al. 1988). Selective allocation of plant-derived C to discrete patches of
fermentation horizon organic matter by ectomycorrhizal mycelia, accompanied by depletion
of N from the organic matter and transfer to the mycorrhizal host plants have also been
demonstrated (Bending & Read 1995). Ectomycorrhizal fungi are thought to have evolved
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repeatedly and independently from saprotrophic precursors, but their genetic potential for
hydrolytic decomposition of cellulose and other plant cell wall components has contracted in
comparison with their saprotrophic ancestors (Kohler et al. 2015; Martin et al. 2016).
However, certain ectomycorrhizal fungal lineages (Bödeker et al. 2009; Bödeker et al. 2014)
appear to have retained the genetic potential to produce the Class II peroxidases (that are
involved in the degradation of recalcitrant polyphenolic compounds in wood and litter-
decaying white-rot fungi) necessary for decomposition of lignin and other phenol-rich
macromolecules (Sinsabaugh 2010), and other species appear to have retained much of the
oxidative decomposition machinery present in brown-rot fungi (Rineau et al. 2012). Some
ectomycorrhizal fungi are thus likely to play a central role in mobilization of N from
recalcitrant humus compounds (Sterkenburg et al. 2018), however, their direct supply of host-
derived sugars means that they are not dependent on organic matter as a source of metabolic
C (Lindahl & Tunlid 2015). This co-metabolic degradation of complex, recalcitrant organic
substrates, facilitated by supply of host-derived sugars to mycorrhizal hyphae releasing N, has
been likened to the priming effect discussed below in relation to release of root exudates into
the rhizosphere (Lindahl & Tunlid 2015). However the direct allocation of carbon to
ectomycorrhizal fungi using different decomposition mechanisms, based on different
combinations of oxidative and hydrolytic enzymes, as well as non-enzymatic Fenton
chemistry (Nicolás et al. 2018), permits efficient extraction of N from a range of organic
substrates. This process is arguably more efficient than rhizosphere priming since the C
allocation is direct and can be controlled by the plant. In addition, because of the extensive
growth of the fungal hyphae, nutrients can be accessed from substrates that are spatially
separated from the roots, without competition from other rhizosphere occupants. Recent work
has demonstrated that some non-mycorrhizal fungi can also transfer N to plants. Thus, plants
can acquire nitrogen from soil insects through their endophytic associations with Metarhizium
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spp.,which are ubiquitous soil-dwelling insect-pathogenic fungi (Behie, Zelisko & Bidochka
2012). Insect-derived nitrogen represented up to 48% of the plant nitrogen content and this
transfer process was driven by reciprocal allocation of C from the plant roots to the fungal
mycelium (Behie et al. 2017). In addition, there is increasing evidence that dark septate
endophytic (DSE) fungi have the potential to facilitate the transfer N to plants (Vergara et al.
2017). DSEs are classified as facultative biotrophs and are known to associate with the roots
of hundreds of plant species. Unlike AM fungi, DSE fungi are saprophytic and have the
potential to break down chemically recalcitrant organic N. Consequently, when the C:N ratio
of organic matter increases and N typically becomes more limiting along successional
gradients, DSEs can outcompete AM fungi leading to greater root nutrient acquisition and
primary productivity (Huusko, Ruotsalainen & Markkola 2017).
A wide range of bacteria have the ability to convert atmospheric N2 into ammonia
through the process of biological N fixation. To benefit from this function, various plants
establish mutualistic interactions by releasing signals in the form of secondary metabolites to
free living N-fixing bacteria so that they can enter roots. The mutualism culminates in the
formation of nodules, a new plant organ on the root. Inside the nodules, the plants supply
photosynthetically fixed C to the bacteria, which in return are committed to provide fixed N to
the host plant. More than 220 plant species belonging to the orders Fagales, Rosales, and
Curbitales (the so-called actinorhizal plants) (Santi, Bogusz & Franche 2013), as well as most
of the 18,000 leguminous plant species, can engage in highly specific mutualistic associations
with soil bacteria. More than a decade ago, it was shown that not only -, but
alsoproteobacteria have coevolved with their legume hosts for up to 50 million years to
establish a N-fixing symbiosis (Moulin et al. 2001; Chen et al. 2003). However, fewer than
10% of the symbionts of the 750 legume genera have been fully characterized and further
studies may reveal additional bacterial taxa that are recruited by plants to meet their N needs.
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In addition to the well-documented N-fixing symbiosis, there is new evidence that plants can
also recruit N2-fixing bacteria that do not form nodules, but rather live in the rhizosphere
(associative) or in the plant tissues (endophytic) (Bouffaud et al. 2016). For instance, under
N-depleted conditions, boreal feather mosses allocate resources to the production of chemo-
attractant for N2 fixing cyanobacteria which then colonize the mosses. In return, these
cyanobacteria contribute to moss nutrition by transferring fixed N2 (Bay et al. 2013). These
symbiotic cyanobacteria can enter epiphytic and intracellular or extracellular endophytic
interactions with their host. In boreal forests, this N2 fixation by cyanobacteria living in
symbiosis with mosses can account to up to 50% of total biological N input (DeLuca et al.
2002; Santi, Bogusz & Franche 2013; Warshan et al. 2017). In N-depleted soils in Mexico,
recent work has shown that the mucilage of aerial roots of an indigenous landrace of maize
enriched in diazotrophs (Van Deynze et al. 2018) (Figure 2). This mucilage, rich in arabinose,
fucose and galactose, can support N2-fixing activity by the diazotrophs. The fixed N2 is
transferred efficiently to the host plant, contributing from 29% to 82% of its N nutrition (Van
Deynze et al. 2018). Altogether, these findings emphasize the need for further research
aiming at identifying overlooked plant traits that can be used to recruit N-fixing bacteria, in
order to help fulfil the plants’ N demand when grown in N-depleted soils.
Recently, there have been considerable advances in understanding the molecular
mechanisms underlying plant-microbe symbiotic associations (Markmann & Parniske 2009;
Oldroyd 2013; Svistoonoff, Hocher & Gherbi 2014; Geurts, Xiao & Reinhold-Hurek 2016;
Floss et al. 2017; Roth & Paszkowski 2017). Although legume-bacteria associations evolved
more than 300 million years later than AM symbiosis, both types of symbiosis begin with a
molecular dialog based on recognition signals between the host plant and microorganisms
with a common set of plant genes involved in signal transduction (Symbiosis Receptor Kinase
- SYMRK) that are shared by all host plants. It has therefore been proposed that the genes
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involved in the symbiosis signaling pathway evolved in the context of the evolutionarily older
AM fungal symbiosis and have later been recruited during the evolution of bacterial root
nodule symbiosis (Kistner & Parniske 2002), including the actinorhizal symbioses in non-
legumes (Venkateshwaran et al. 2013). Gaining more detailed information on this symbiosis
signaling pathway may assist in better understanding general plant strategies related to
mutualistic association with microorganisms to acquire nutrients.
Having developed strategies to acquire nutrients through mutualistic interactions,
plants must continue to evolve to maintain beneficial symbionts and reduce the fitness
benefits from microbial cheater, i.e. symbionts providing little benefit to their host plants. A
principal problem for plants is that cheaters can mimic the signals of their cooperative
competitors in order to gain access to a plant host and therefore recognition signals are not
reliable for selecting the best microbial partners (van’t Padje, Whiteside & Kiers 2016). Plants
have therefore evolved mechanisms to enforce cooperation by sanctioning symbionts that are
inefficient suppliers of nutrients (Kiers & Denison 2008; Nehls 2008). For example, Kiers et
al. (2011) demonstrated that cooperation in mycorrhizal symbiosis is stabilized by reciprocal
rewards involving increased supply of carbohydrates and nutrients.
What is evident from the examples highlighted above, is that plants can adapt to N-
limiting environments not only by recruiting microorganisms through mutualistic interactions
but also by stabilizing the effectiveness of these interactions. This supports the idea that plants
are not standalone entities; rather, they should be considered from a more holistic perspective,
as holobionts, including the full diversity of the many different microorganisms associated
with them (Vandenkoornhuyse et al. 2015).
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Priming soil organic matter decomposition by microorganisms
Evidence is growing to suggest that rhizosphere priming represents an important strategy for
plants to access organic N in soil. Here we define priming as a short-term change in the
mineralization of native soil organic matter (SOM) caused by the release of rhizodeposited
carbon (C) from live roots (including root exudates, sloughed cells, epidermal cell death,
etc…). This release of labile C often causes the rhizosphere microbial community to grow
resulting in an increased microbial demand for N. While this N can be supplied by low
molecular weight root exudates (e.g. amino acids), if the C/N ratio of the rhizodeposits is high
then the microbial N demand is typically satisfied by releasing proteases and deaminases to
release N from N containing polymers held in SOM (e.g. protein, chitin). Most research to
date on priming has focused on its role in soil C dynamics, with studies showing that it can
both enhance and suppress SOM mineralization (Jones, Hodge & Kuzyakov 2004; Kuzyakov
2010; Cheng et al. 2014; Huo, Luo & Cheng 2017). Plant traits such as root life span and
nutrient stoichiometry can drive variation in soil N/P, which has been linked to SOM
decomposition (Carrillo et al. 2017). Priming effects on SOM are therefore considered to be
of high importance for the global C cycle due to their influence on C fluxes from soil
(Heimann & Reichstein 2008; Cheng et al. 2014). However, recent studies also suggest that
rhizosphere priming can influence soil N cycling and N supply to plants (Dijkstra et al. 2013;
Zhu et al. 2014; Meier, Finzi & Phillips 2017), and that it might even be an evolutionary
strategy developed between plants and rhizosphere microbes (including symbiotic
mycorrhizal fungi) that benefit from plant-derived C and, in turn, benefit their plant hosts
through enhanced microbial mineralization of soil organic N (Cheng et al. 2014; Mwafulirwa
et al. 2017). Furthermore, effects of priming appear to be especially important for plant N
supply under global change, with studies demonstrating that priming effects on soil N cycling
- driven by increased rhizodeposition and stimulation of extracellular enzymes involved in N
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mineralization - are a key mechanism by which trees sustain long-term increases in growth
under elevated CO2 (Langley et al. 2009; Phillips, Finzi & Bernhardt 2011).
Reports on the magnitude and direction of priming effects on soil-N cycling vary
considerably and many factors have been shown to contribute to such variation, including
plant species identity and phenology, mycorrhizal status, and the type of soil (Dijkstra et al.
2009; Bengtson, Barker & Grayston 2012; Zhu et al. 2014; Mwafulirwa et al. 2017).
However, one key determinant of priming is soil nutrient availability, especially the extent to
which the soil and the microbial community are N limited. Various hypotheses have been put
forward explaining the link between priming effects and nutrient status of soil, with the
general view being that it is of greater significance in N-limited soils (Dijkstra et al. 2013;
Cheng et al. 2014). One hypothesis, for example, is that in N-limited soils, root exudates
cause positive priming due to the stimulation of microbial extracellular enzymes (i.e.
proteases) involved in the breakdown of soil organic N pools, which serves to meet microbial
N demand as well as enhancing microbial N mineralization and N supply to plants (Phillips,
Finzi & Bernhardt 2011; Zhu et al. 2014; Meier, Finzi & Phillips 2017). However, plant and
microbial N demand can be high in low N soils, and if plants compete effectively with
microbes for N this can result in negative priming due to a reduction in microbial
decomposition (Dijkstra et al. 2010). Negative priming is also thought to occur in N replete
soils, where microbes may use exudates to meet their C and energy needs, rather than mining
SOM for N, thereby leading to reduced mineralization of organic N (Cheng 1999; Dijkstra et
al. 2013). Further, in soils that are limited more by phosphorus than N, it has been suggested
that exudates might be used to mobilize phosphorus, rather than mine SOM for N, thereby
reducing N mineralization and supply to plants (Dijkstra et al. 2013). It should be noted,
however, that measurement of N priming is far more challenging than for C priming due to
many unresolved methodological challenges.
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While it is clear that rhizosphere priming can influence mineral N availability in soil,
uncertainties remain about the mechanisms by which root exudation affects microbial
mineralization-immobilization dynamics and hence the supply of N to plants, as well as its
effects on subsequent microbial N-cycling processes. One major uncertainty concerns the role
of microbial community composition in determining the outcome of priming effects for plant
N supply, especially given that microorganisms and different microbial groups vary in their
functional capabilities regarding the decomposition of SOM (Chen et al. 2014). Another
source of uncertainty concerns the role of plant traits (Violle et al. 2007) in influencing
rhizosphere priming. In particular, very little is known about how variations in the amount
and quality of exudates released from plants influence priming effects (Carrillo et al. 2017),
and how they are influenced by plant N status, soil abiotic properties, such as soil moisture,
texture and pH, and the formation of aggregates which stabilize and protect organic matter
from microbial mineralization (Cheng et al. 2014). Despite early studies highlighting the role
of faunal-microbial interactions in the rhizosphere for the mineralization and plant uptake of
soil N (Clarholm 1985; Ingham et al. 1985; Ritz & Griffiths 1987), major uncertainties exist
about the influence of plant exudates on nutrient release via these trophic interactions (Cole et
al. 2004; Blagodatskaya et al. 2014). Our poor ability to measure the spectrum and amounts
of root exudates in soil in a quantitatively rigorous manner (Oburger & Jones 2018) calls for
further research to overcome this bottleneck. In the case of priming, it is highly likely that
both negative and positive priming will occur in different regions of the root system (due to
spatial patterns of root exudation, microbial activity and soil organic C and N heterogeneity).
Consequently, more studies on the spatial and temporal patterning of priming are required.
Finally, considering that N-processes are inextricably interlinked, studies explicitly
integrating downstream N-cycling processes using mineral N as substrates when focusing on
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N-priming in the rhizosphere are needed to fully account for the role of plant traits on the
entire N cycle and fill an important knowledge gap.
Plants limit N losses to conserve available N
Nitrification and denitrification are the microbial processes primarily responsible for losses of
mineral N from terrestrial ecosystems. During denitrification, oxidized forms of N - nitrate
and nitrite - are used as electron acceptors by microorganisms for respiration when oxygen is
limiting and are successively reduced into NO, N2O and N2 gases, which are returned to the
atmosphere. Nitrification consists in the oxidation of ammonium into nitrate, which can be
leached into groundwater or converted into N gases by denitrification, thereby causing N
losses from the soil-plant system (Philippot et al. 2009). In the following section, we highlight
how plants can limit these microbial processes leading to N-losses through a range of
mechanisms and traits.
Recent studies indicate that relationships exist between plant growth strategies and/or
plant traits, the activity of N-cycling microbes, and N retention and loss. Based on the
analysis of a range of plant species it was demonstrated that, in comparison to conservative
species, species with exploitative growth strategies are associated with reduced N losses via
leaching (de Vries & Bardgett 2016) and reduced microbial N2O emissions (Abalos, van
Groenigen & De Deyn 2018). Furthermore, Cantarel et al. (2015) showed that nitrification
rates are positively related to specific root length (i.e. root length per unit of root biomass),
root N as well as plant affinity for NH4+ while root length density (root length per unit of soil
volume) was identified as the key trait regulating the effects of plants on N2O emissions
(Abalos et al. 2014; Cantarel et al. 2015; Abalos, van Groenigen & De Deyn 2018). Plant
traits related to acquisitive strategy, e.g. high specific root length (Roumet et al. 2016) and/or
high root length density, are commonly associated with high rates of soil-resource acquisition
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(Wright et al. 2004; Roumet, Urcelay & Díaz 2006; Abalos, van Groenigen & De Deyn
2018). Therefore one interpretation for these linkages between plant traits and the activity of
N-cycling microbes is that plants and microbes compete for N. Indeed, Moreau et al. (2015)
found that the abundance of bacteria capable of using NO3- as an alternative electron acceptor
when oxygen is limiting, was negatively related to root N uptake rate. This indicates that
microorganisms performing the first step of denitrification can be outcompeted by plants with
a high root N uptake rate. However, previous studies of competitive interactions between
plants and microbes have mostly focused on N-assimilatory processes (Kuzyakov & Xu 2013)
whereas microbial dissimilatory processes using N to obtain energy have been overlooked.
In addition to the above-mentioned exploitative competition, plants have also evolved
direct interference competition mechanisms (i.e. allelopathy) to conserve N by producing
secondary metabolites detrimental to microorganisms that cause N losses from soil. An
inhibition phenomenon of N-cycling microorganisms by plants was first evidenced by the
finding of lower nitrification rates in tropical grassland ecosystems (Lata et al. 2004).
Following this, it was shown that nitrification inhibitors can lead to a decline of up to 90% in
ammonia oxidation rates in Brachiara pasture and a lower abundance of both archaeal and
bacterial ammonia-oxidizing microorganisms (Subbarao et al. 2009). Such a decrease in the
abundance of ammonia oxidizers on plant roots compared to the bulk soil has also been
reported for two forage crops, Medicago sativa and Dactylis glomerata (Zhao et al. 2017).
These results have been nicely complemented by laboratory-based studies, which
demonstrated the presence of secondary metabolites released into soil via root exudates, such
as methyl 3-(4 hydrophenyl) propionate or cyclic diterpene, that block the ammonia-
monooxygenase and hydrolamine oxidoreductase in the nitrification pathway, a phenomenon
called biological nitrification inhibition (BNI) (Zakir et al. 2008; Subbarao et al. 2009). The
release of these inhibitors is not a passive process, but rather is triggered by high NH4+
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concentration in the root environment (Subbarao et al. 2007b; Zakir et al. 2008). This was
evidenced by using a split-root system in which half of the root system was exposed to NH4+
and the other half to NO3-. Only the part of the root system exposed to NH4
+ triggered the
release of the brachialactone, indicating a localized active process. Presence of the BNI trait
in plants is now being identified in a growing number of species (Subbarao et al. 2007a;
O’Sullivan et al. 2016), emphasizing its potential to increase nitrogen use efficiency in
agriculture to conserve N in the NH4+ form (Coskun et al. 2017). Another groundbreaking
finding was the demonstration that some plants can also inhibit denitrification by up to 80%
through the release of procyanidins in root exudates (Bardon et al. 2014; Bardon et al. 2016).
In contrast to nitrification inhibitors, the impact of such denitrification inhibitors has not yet
been quantified in the field. This ability to control N-cycling microbial communities through
inhibition has been proposed as a mechanism allowing invasive plants to outcompete native
plant species, yet empirical tests remain rare (Hawkes et al. 2005; Dassonville et al. 2011;
Boudsocq et al. 2012; Yelenik & D'Antonio 2013; McLeod et al. 2016). On the other hand, it
is important to note that plants can also stimulate denitrification, for example via the release
of organic compounds by rhizodeposition, as described above (Højberg, Binnerup & Sørensen
1996; Henry et al. 2008). Thus, the nature of root exudates can differentially influence
denitrifiers and N losses (Henry et al. 2008; Guyonnet et al. 2017) but the positive
rhizosphere effect on denitrification is confined to air-filled porosities lower than 10-12%
(Prade & Trolldenier 1988). Nevertheless, the exact factors behind the stimulation of
denitrification in the vicinity of plant roots are still debated since most of them are strongly
interwoven.
To sum up, plant traits that can negatively affect microbial communities responsible
for N losses from the soil-plant system are numerous. Such plant traits are not only involved
in exploitative competition, i.e. restricting the supply of N to microorganisms, but also in
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direct interference competition i.e. by producing inhibitory secondary metabolites to harm
microorganisms. However, further studies are needed in order to determine how widespread
this interference competition strategy is as a means for plants to control microbial guilds
involved in N transformation and losses. It is also worth considering that traits which are not
involved in competitive interactions between plants and soil microorganisms can decrease N-
losses. Indeed, it has recently been reported that mutualistic relationships between plants and
AM fungi reduce N2O losses from soil (Bender et al. 2014), yet the underlying mechanisms
and traits remain unclear. Another process that can help to limit N losses is from soil is the
respiratory reduction of nitrate to the more stable form ammonium (DNRA) by bacteria and
fungi, which is more widespread in soils than previously thought (Rütting et al. 2011). We
argue that, due to the higher C in the root vicinity, DNRA is likely favored in the rhizosphere,
allowing plants to conserve N in soil. This is supported by a recent study showing that
bacteria possessing a key gene involved in DNRA were overrepresented in the rhizosphere
compared to bulk soil (Li et al. 2014).
Conclusions and future directions
Global cycling of N has transgressed a critical threshold: the amount of atmospheric N2
industrially converted into ammonia to produce fertilizers now exceeds that produced from all
of the Earth’s terrestrial processes (Rockström et al. 2009; Canfield, Glazer & Falkowski
2010). We show here that plants are not passive conduits, taking up whatever N diffuses to
their roots. Rather they can influence both the availability and uptake of different forms of N
through their interactions with various nitrogen cycling microbial guilds. This influence of
plants on N-cycling microorganisms is not limited to a few N-cycling processes as previously
assumed (Figure 3). Indeed, we highlight the fact that plants can improve their N nutrition by:
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(i) establishing various types of symbiosis with soil microorganisms; (ii) stimulating the
activity of microorganisms in the root vicinity to increase N availability; (iii) increasing N
conservation in soil by limiting microbial processes that lead to N losses, such as nitrification
and denitrification, directly through the release of inhibitors from their roots. Plants can also
adversely affect N-cycling microbes indirectly, through competition for N, with higher plant
N uptake rates decreasing soil N availability with consequences for the abundance and/or
activity of microbes.
Many mechanisms exist by which plants can modify N cycling processes mediated by
microbes; as such, we suggest that a plant-trait based framework could serve to advance our
understanding of linkages between plant N acquisition strategies and the activities of N-
cycling microbial guilds (Figure 3). As is evident from our review, the capacity of plants to
influence different N-cycling microorganisms is often broadly linked to their resource
acquisition strategy, and whether they have traits related to resource exploitation or
conservation. However, many uncertainties exist, especially concerning the mechanistic
nature of relationships between root traits and N cycling-microorganisms, and possible
cascading effects within the N-cycle given that N cycling processes are inextricably linked.
We also argue that studying plants from N-poor habitats may assist in better understanding
plant traits directly controlling N-cycling microorganisms. Indeed, if we are to transition from
high input systems to low input systems and exploit microbial properties that might be
valuable in low-input sustainable systems, we need more information about those systems –
not just conventional agricultural systems where large amount of mineral N inputs are
commonly used. In addition, significant work is still required to better understand the spatial
and temporal dynamics of N turnover and root N uptake in the rhizosphere and its association
with plant traits (e.g. root morphology, mycorrhizas). It is already established that N cycling
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hotspots exist in soil (Hill & Jones 2019), however, the extent to which roots create or
capitalize on these ephemeral hot spots remains largely unexplored.
This influence of plants on N-cycling microbes might also be related to the interplay
between microorganisms and microbial-feeding microfauna (“the microbial loop”), for
example through the release of root exudates. While there is little evidence for the notion that
microbial feeding soil microarthropods are regulators of microbial–plant competition for N
(Cole et al. 2004), it has been shown that large bodied fauna can magnify the effects of litter
composition on N mineralization (Carrillo et al. 2011). However, such interactions have been
overlooked and further studies are needed to understand the contribution of plant-microbe-
microfauna interactions in soil N-cycling. There is also a need for improved knowledge of the
reciprocal capacity of N-cycling microbes to shape plant function and community structure,
the so-called plant soil feedback (Van der Putten et al. 2013). Furthermore, a new area of
study could emerge by addressing the lack of knowledge related to how N cycling in the
rhizosphere can affect plant health. Fluxes of NO from soil during N transformation have
been quantified in several studies for their role in atmospheric chemistry (Medinets et al.
2015). However, NO is also an important messenger in plant signaling and defense (Besson-
Bard, Pugin & Wendehenne 2008) and, despite median NO flux of 3.2 ng N m−2 s−1 (Huang &
Li 2014), the possible role of NO emitted in the rhizosphere by microbes in plant disease
resistance has not yet been considered.
Authors' Contributions
All authors contributed to the writing the manuscript.
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Figure 1: Factors affecting the forms of plant nitrogen uptake. The proportion of the
different N forms that plants absorb depends on many interacting factors. Even though the
mechanisms underlying some of these factors are still elusive and interactions among factors
are not well understood, the relative amount of N in soils is identified as the main factor
determining which forms of N are used by plants (von Wirén, Gazzarrini & Frommer 1997;
Chapin, Matson & Mooney 2002). Soil pH, moisture and temperature all influence the forms
of N in soil and, therefore, the forms of N uptake by plants (Britto & Kronzucker 2013). Soil
potassium can also be a determinant due to its ability to alleviate ammonium toxicity in nitrate
specialist plants (Li et al. 2012). Competing neighbouring plants (Miller, Bowman & Suding
2007) and competing soil microorganisms (Kuzyakov & Xu 2013) may also reduce the
availability of some N forms, while symbiosis can enhance the capacity of the plants to
absorb certain forms of N. Plant properties are the other key factor affecting the forms of N
uptake by plants. The plant species/genotype act mainly via characteristics linked to the
transporters of the different N forms (von Wirén, Gazzarrini & Frommer 1997) and
susceptibility to ammonium toxicity (Britto & Kronzucker 2002), while the plant N status and
plant growth stage (Cui et al. 2017) can differently affect nitrate and ammonium uptake rates.
Accordingly, some species are specialised while others possess an important plasticity in the
form of N uptake (Britto & Kronzucker 2002; Houlton et al. 2007). These differences have
been suggested to affect interactions with soil microorganisms both in terms of competition
and cooperation (Britto & Kronzucker 2013).
Figure 2: Aerial root mucilage from an indigenous landrace of maize growing in N
depleted soil (reproduced from Van Deynz et al. 2018).
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Figure 3: A schematic of how plants can influence N-cycling processes to increase N
availability. Plant traits affecting the N-cycling processes are shown in blue; some traits are
more related to N-cycling processes (in boxes) while others affect all the processes (top of the
figure).
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