mechanisms and regulations of priming effects in soil li, jian

LUND UNIVERSITY

PO Box 117221 00 Lund+46 46-222 00 00

Mechanisms and regulations of priming effects in soil

Li, Jian

2019

Link to publication

Citation for published version (APA):Li, J. (2019). Mechanisms and regulations of priming effects in soil. Department of Biology, Lund University.

Total number of authors:1

General rightsUnless other specific re-use rights are stated the following general rights apply:Copyright and moral rights for the publications made accessible in the public portal are retained by the authorsand/or other copyright owners and it is a condition of accessing publications that users recognise and abide by thelegal requirements associated with these rights. • Users may download and print one copy of any publication from the public portal for the purpose of private studyor research. • You may not further distribute the material or use it for any profit-making activity or commercial gain • You may freely distribute the URL identifying the publication in the public portal

Read more about Creative commons licenses: https://creativecommons.org/licenses/Take down policyIf you believe that this document breaches copyright please contact us providing details, and we will removeaccess to the work immediately and investigate your claim.

https://portal.research.lu.se/en/publications/c1714c60-147b-4fca-ac83-57ef7e191abf

JIAN

LI

Mechanism

s and regulations of priming effects in soil

2019

Department of BiologyFaculty of Science

Lund University, ISBN 978-91-7895-360-8

Mechanisms and regulations of priming effects in soilJIAN LI

FACULTY OF SCIENCE | LUND UNIVERSITY

List of papers

I. Li, J., Bengtson, P. Quantification of the rhizosphere priming effect differs widely depending on the choice of control

II. Li, J., Zhou, M., Alaei, S., Bengtson, P. Rhizosphere priming effects differ between Norway spruce (Picea abies) and Scots pine (Pinus sylvestris) seedlings cultivated under two levels of light intensity.

III. Li, J., Alaei, S., Zhou, M., Bengtson, P. Root influence on soil nitrogen availability and microbial community dynamics results in contrasting rhizosphere priming effect in pine and spruce soil.

IV. Alaei, S., Li, J., Jackowica-Korczynski, M., Bengtson, P. The influence of elevated CO2 and N fertilization on rhizosphere priming, C sequestration and N cycling in soil planted with Picea abies seedlings.

V. Wild, B., Li, J., Pihlblad, J., Bengtson, P., Rütting, T. 2019. Decoupling of priming and microbial N mining during a short-term soil incubation. Soil Biology and Biochemistry 129, 71-79.

9789178

953608

1


Jian Li

DOCTORAL DISSERTATION

By due permission of the Faculty of Science, Lund University, Sweden. To be defended at Blue Hall, Ecology Building, Lund, Sweden on 13th of Dec

2019 at 10:00

Faculty opponent Dr. Feike Dijkstra

The University of Sydney, Sydney, Australia

2

Organization LUND UNIVERSITY

Document name Doctoral dissertation

Department of Biology, Ecology Building, SE-223 62 Lund, Sweden

Date of issue 2019-12-13

Author: Jian Li Sponsoring organization

Title and subtitle: Mechanisms and regulations of priming effects in soil

Abstract Soil organic matter (SOM) plays a significant role in the global carbon (C) and nutrients cycles, not only since SOM is a major storage reservoir of C and plant nutrients, but also since microbial decomposition of SOM releases large quantities of CO2 to the atmosphere. The accelerated or reduced turnover rate of SOM after the input of labile C compounds is a phenomenon known as the priming effect (PE). When the labile C originates from root exudate, the effect is referred to as the rhizosphere PE. The rhizosphere PE has been reported to increase SOM decomposition by up to 380%, but also to reduce SOM decomposition by up to 50% compared to rootless soil. Although PE is an important regulator of SOM decomposition and nutrient cycling, little is known about the mechanisms and regulations of PEs. In this thesis, I performed several greenhouse experiments with living tree seedlings to investigate how different tree species, light intensity, root exudation rate, elevated CO2 and nitrogen (N) fertilization influence the rhizosphere PE on SOM decomposition and gross N mineralization. I also performed an incubation experiment to compare the effect of increased C and N availability on the depolymerization of native SOM. I further measured the microbial growth rate, microbial community composition and potential enzyme activities to investigate the underlying mechanisms and role of different microbial groups in regulating the PE. My results showed that the magnitude and direction of the rhizosphere PE differed between plant species. Light intensity also played an important role in regulating the rhizosphere PE, depending on the plant species present. My results further showed that the effect of N-fertilization and elevated CO2 on the partitioning of N between plants and microbes was of importance for SOM decomposition and PEs. Moreover, rhizosphere PE was not dependent on the root exudation rate. Instead, my studies suggest that variation in soil N availability is a key regulator of the PEs. My studies further revealed that fungi are responsible for the PEs, while potential enzyme activities are poor indicators of the PEs. In conclusion, my studies suggest that soil C and N cycling are strongly affected by the PEs caused by living roots. Therefore, the regulation of PEs on SOM decomposition and N mineralization could play an important role in the terrestrial ecosystem's feedback to global change in response to e.g. elevated CO2 and N deposition. My results also demonstrated that the PE is regulated by multiple environmental factors, e.g. plant species, light intensity, atmospheric CO2 concentration, and soil N availability. In order to accurately predict how the PE on soil C and N cycling responds to environmental change, these factors need to be considered more in future studies.

Keywords: Soil organic matter decomposition, gross N mineralization, root exudate, plant species, soil C: N, microbial growth rate, microbial community composition, enzyme activity Classification system and/or index terms (if any)

Supplementary bibliographical information Language: English

ISSN and key title ISBN 978-91-7895-360-8

Recipient’s notes Number of pages 56 Price

Security classification

I, the undersigned, being the copyright owner of the abstract of the above-mentioned dissertation, hereby grant to all reference sources permission to publish and disseminate the abstract of the above-mentioned dissertation. Signature Date 2019-11-04

3


Jian Li

4

Cover design: Jian Li

Copyright: Jian Li

Faculty of Science Department of Biology ISBN 978-91-7895-360-8 (print) 978-91-7895-361-5 (pdf) Printed in Sweden by Media-Tryck, Lund University Lund 2019

5

面朝大海

春暖花开

6

Table of Contents

List of Papers .................................................................................................. 8 Author Contributions ..................................................................................... 9 Popular science summary ............................................................................. 10

Introduction .......................................................................................................... 11 Definition of the priming effect ................................................................... 11 Mechanisms .................................................................................................. 13

Positive priming effects ....................................................................... 13 Negative priming effects ..................................................................... 14

Regulating factors ........................................................................................ 17 Microbial activity, biomass and community composition ................... 17

Microbial community composition ............................................................... 17 Plant species and physiology ............................................................... 18 Soil properties ...................................................................................... 19 Amount of the ‘primer’ ....................................................................... 22

Aim and approaches of studies ............................................................................ 23 Paper I .......................................................................................................... 23 Paper II ......................................................................................................... 24 Paper III ........................................................................................................ 25 Paper IV ....................................................................................................... 26 Paper V ......................................................................................................... 27

Methodology .......................................................................................................... 29 Priming of N mineralization ......................................................................... 29 Priming of SOM decomposition .................................................................. 30 Root exudation rate ...................................................................................... 31 Microbial growth rate ................................................................................... 32 Microbial community composition .............................................................. 33 Potential enzyme activities ........................................................................... 33

7

Main findings ........................................................................................................ 35 Evaluation of different controls .................................................................... 35 Plant species effects on rhizosphere PEs ...................................................... 36 Light intensity effect on rhizosphere PEs..................................................... 38 Elevated CO2 and N fertilization effects on rhizosphere PEs ...................... 39 Root exudate effects on the rhizosphere PE ................................................. 39 Soil C and N availability effect on the depolymerization ............................ 41 The role of the different microbial groups in the PEs .................................. 41 The role of the extracellular enzyme in the PEs ........................................... 42

Conclusion and future perspectives .................................................................... 43

Acknowledgments ................................................................................................. 45

References ............................................................................................................. 47

8

List of Papers In this thesis, the papers are referred to by the following roman numerals:






Paper V is reprinted with the permission from the publisher.

9

Author Contributions I. JL, and PB designed the experiment. JL performed the experiment with help

from MZ and SA. JL and PB wrote the manuscript. II. JL, PB, SA and MZ designed the experiment. JL, MZ and SA performed the

experiment. JL and MZ analyzed the data. JL and MZ wrote the manuscript with assistance from PB and SA.

III. JL, PB and SA designed the experiment. JL, MZ and SA performed the experiment. JL analyzed the data and wrote the manuscript with assistance from PB. All co-authors provided comments on manuscript drafts.

IV. JL, PB and SA designed the experiment. JL, MJ and SA performed the experiment. SA analyzed the data and wrote the first draft of the manuscript, with assistance from JL and PB. All co-authors provided comments on manuscript drafts.

V. BW, TR and PB designed the experiment. JL, BW and JP performed the experiment, JL and BW analyzed the data. BW wrote the manuscript, with the assistance of the co-authors.

JL: Jian Li

PB: Per Bengtson

SA: Saeed Alaei

MZ: Moyan Zhou

BW: Birgit Wild

JP: Johanna Pihlblad

TR: Tobias Rütting

MJ: Marcin Jackowica-Korczynski

10

Popular science summary Soils contain more organic carbon (C) than the total terrestrial biomass combined. A large part of this C is stored in complex soil organic matter (SOM) with long turnover times. SOM is also the dominant pool of nutrients required for plant and microbial growth, but the availability of these nutrients is dependent on microbial decomposition of SOM into bioavailable forms. Labile C compounds exuded by plant roots can regulate the decomposition rate of SOM. This belongs to the phenomena of priming effects (PEs). Altered rates of native SOM turnover in the presence of labile C compound is increasingly considered as a mechanism central for the coupling of plant and soil biological productivity. The development of high precision isotopic techniques now enables us to directly quantify PEs, and ample evidence suggests that the PEs are large enough to influence the C balance of terrestrial ecosystems, with implications for the global C cycle and atmospheric CO2 concentration.

Plant's roots can exude labile C compounds into the soil. Accordingly, living roots can increase SOM decomposition by up to 380% or reduce it by up to 50% compare to rootless soil, through the PEs. This large variation can be caused by many controlling factors. For instance, plant species, root exudates, soil N availability, microbial activity, and community composition have been shown to significantly affect PE. Accordingly, several potential mechanisms have been proposed to explain why PE occurs. For instance, positive PE can be caused by microbial activation in response to the input of labile C compounds. In addition, when N is not sufficient for microbial growth, microorganisms will use labile C to mineralize N out of SOM, resulting in stimulation of SOM decomposition and positive PEs. Preferential utilization of the labile C compounds by soil microorganisms, meaning that they do not have to decompose SOM to gain C, have been suggested as a potential reason for negative PEs. Another proposed explanation to negative PEs is that low N availability results in competition for N between plants and microorganisms, leading to a reduction of microbial ability to decompose SOM. Each of these mechanisms may act alone or together, further complicating our understanding of why both positive and negative PEs can occur.

In this thesis, I aimed to investigate the regulation of PEs in soil, with a particular focus on how the input of labile C (e.g. root exudates and sugar) influence SOM decomposition and N mineralization. In summary, my thesis suggested that soil C and N cycling are strongly affected by PE. The PEs were interactively controlled by multiple regulators, including the plant species, light intensity, CO2 concentration, and N availability. My finding also suggested that variation in soil N availability was of particular importance for regulating the PEs, and different subsets of the microbial community are responsible for inducing positive or negative PEs. Taken together, my results suggest that PEs on SOM decomposition lay an important role in the ecosystem's feedback to global change in response to e.g. elevated CO2 and N deposition.

11

Introduction

The largest terrestrial carbon (C) pool is stored in soils as complex soil organic matter (SOM) (Yiqi and Zhou, 2010; Lal, 2018), and at the global scale, nearly half of total soil respiration originates from the decomposition of SOM (Davidson and Janssens, 2006). As a result, even a small change in the SOM decomposition rate can have a strong influence on the atmospheric CO2 concentrations, and hence the global climate (Davidson and Janssens, 2006; Heimann and Reichstein, 2008). The decomposition of SOM is dependent on transformations mediated by soil microorganisms that decompose SOM to gain energy and nutrients. The decomposition activities of these microorganisms are in turn strongly affected by the presence of living roots, since their biomass and activity are promoted by labile C compound supplied in root exudates (Van Hees et al., 2005). In fact, changes in decomposition of native SOM in response to input of labile C compounds, referred to as priming effects (PEs), is increasingly considered a central mechanism for soil C and nutrient turnover, which potentially plays a key role in regulating the global C cycle (Cheng et al., 2014; Guenet et al., 2018).

Definition of the priming effect The PE was first identified by Lohnis in 1926. It is a phenomenon where the SOM decomposition rate is changed after the addition of exogenous labile C substrates (e.g. root exudates, fresh plant residues, manure, dead microorganisms). The SOM decomposition rate can be increased or decreased after the input of the labile C, referred to as positive PE and negative PE, respectively (Kuzyakov, 2002; Cheng et al., 2014) (Fig. 1). Positive PEs can strongly influence SOM decomposition rates, and large amounts of C and N could be mobilized in a very short time as a result (Kuzyakov et al., 2000; Zhu et al., 2014).

12

Fig. 1. The priming effect (PE) - interactions between added labile C compound and soil organic matter (SOM) decomposition: (a) acceleration of SOM decomposition - positive PE; (b) reduction of SOM decomposition - negative PE. (Adapted from Kuzyakov 2002).

PEs are divided into two classes: real PE and apparent PE. Jenkinson et al. (1985) suggested that real PE is an increase in the decomposition of recalcitrant SOM, whereas apparent PE is an increase of microbial C turnover, which is not linked to changes in SOM decomposition. In other words, real PEs are a change in the SOM decomposition rate by input of labile C compound, while apparent PE results from an acceleration of the CO2 evolution in response to the activation of microbial metabolism and higher microbial biomass turnover, which is not related to SOM turnover (Blagodatskaya and Kuzyakov, 2008).

In this thesis, I determined PEs induced by the input of labile C compounds such as sugar, and also the rhizosphere PEs, which is the change in SOM decomposition and gross N mineralization caused by labile C exuded by living roots.

13

Mechanisms Several possible mechanisms of the PEs have been proposed in recent years. In the following section, I summarized the most commonly proposed mechanisms.

Positive priming effects

Microbial activation Soil microbes have a starving-survival lifestyle of dormancy or low activities (Hobbie and Hobbie, 2013). The microbial activation theory suggests that the input of glucose or other labile C compounds can trigger microorganisms to an active stage, leading to an increase of microbial activity, biomass and SOM decomposition (Kuzyakov, 2002; Cheng and Kuzyakov, 2005). Accordingly, several studies have demonstrated that stimulated SOM decomposition in response to labile C input is accompanied by higher microbial activity and biomass (Xiao et al., 2015; Kumar et al., 2016; Li et al., 2018). Even a very low amount of labile C can activate microorganisms and increase the SOM decomposition rate (Kuzyakov et al., 2000; Cheng et al., 2014), which might explain why PEs often result in a disproportional increase in SOM decomposition relative to the amount of labile C that is added or exuded by living roots. For instance, Bengtson et al. (2012) found each mg C exuded by roots C released 6 mg bioavailable C from SOM.

Increased enzyme production and activity SOM decomposition is mediated by extracellular enzymes secreted by soil microorganisms (Marx et al., 2001; Sinsabaugh et al., 2008). The input of labile C substrates can serve as an energy source for the synthesis of such enzymes, resulting in an increase of SOM decomposition (Blagodatskaya and Kuzyakov, 2008). Accordingly, changes in extracellular enzyme production, promoted by the addition of labile C, have been suggested as a possible explanation for PEs. Labile C input might not only increase the microbial production of enzymes, but also stimulate the activity of enzymes already present in the soil. Oxidative enzymes that depolymerize macromolecules and produce soluble substrates for microbial assimilation appear to be particular importance for the initial stages of SOM decomposition (Hofrichter et al., 2010). Root exudates can stimulate the activity of these enzymes, by providing the substrate for accessory H2O2-generating oxidases (Ander and Marzullo, 1997). Accordingly, living roots have been found to stimulate the potential activity of both hydrolytic and oxidative enzymes (Zhu et al., 2014; Kumar et al., 2016). However, there are also examples of positive PE that did not co-occur with an increase in potential enzyme activity (Wild et al., 2017). These ambiguous results demonstrate that there is no straightforward relationship between

14

enzyme activity and PEs, possibly since the methods used measure the concentration of enzymes, and not the in situ enzyme activity (Tiedje, 1982).

Microbial N-mining The N-mining hypothesis suggests that when inorganic N is limiting microbial growth and activity, microbes in the rhizosphere utilize C-rich rhizodeposits to ‘mine’ N out of SOM (i.e. mineralize organic N or depolymerize proteins and other organic N sources), which results in an accelerated SOM decomposition and positive PEs (Kuzyakov et al., 2000). This can explain why the input of easily available C often increases the CO2 efflux from native soil if microorganisms become N-limited when supplied with an abundant source of C and start to actively degrade organic materials to gain N (Garcia-Pausas and Paterson, 2011). In contrast, when inorganic N is sufficient, microbes in the rhizosphere assimilate the C exuded by roots instead of decomposing SOM to release available C, which results in lower SOM decomposition and a negative PEs (Preferential substrate utilization, see below) (Cheng, 1999). The N mining hypothesis is supported by several studies that have found a close connection between soil N content and the PE (Chen et al., 2014; Qiao et al., 2016; Xu et al., 2018).

Negative priming effects

Preferential substrate utilization The input of labile C into the soil does not always result in increased SOM decomposition. A recent review paper showed that the PE caused by the labile C input can also reduce the decomposition of SOM, with as much as 50% (Cheng et al., 2014). This can be explained by the preferential substrate utilization hypothesis, which suggests that microorganisms use the most energy-rich and easily available substrates first (Cheng, 1999). Therefore, under conditions where mineral nutrients are sufficient, the microorganisms are more likely to use labile root-derived C than more complex SOM-derived C, leading to decreased decomposition of SOM. Accordingly, some studies have found that the SOM decomposition rate decreased in N fertilization treatments relative to unfertilized controls (He et al., 2016; Aye et al., 2018; Hicks et al., 2018).

Plant-microbial competition Competition between roots and soil microorganisms for available N may suppress microbial growth and activity when N availability is limited (Månsson et al., 2009) and subsequently the SOM decomposition rate (Kuzyakov, 2002; Kuzyakov and Xu, 2013). This mechanism has been suggested as a possible cause of the negative PEs since the input of labile C in root exudates might aggravate the competition (Kuzyakov and Xu, 2013). Accordingly, it has been observed that the PEs change

15

from negative to positive with N fertilization (Chen et al., 2014; Lu et al., 2018), but there are also observations on the contrary (Aye et al., 2018; Hicks et al., 2018). It seems like the competition mechanism dominates under extremely N limiting conditions, while the N-mining mechanism will be dominant when N is at a moderately limited level, and the preferential substrate use mechanism will dominate when mineral N is abundant (Kuzyakov et al., 2000).

In reality, these mechanisms may individually or in combination regulate the PEs. Which mechanism that dominates could depend on the specific circumstances since many abiotic and biotic factors (e.g. soil physicochemical properties, soil microbial activity and community composition, plant community structure, etc.) can influence the direction and magnitude of PEs.

17

Regulating factors As mentioned above the magnitude and direction of PEs vary considerably among studies (Cheng et al., 2014). The large variation can be caused by many factors. Variations in microbial community composition and activity can have a profound effect on SOM decomposition rates (Strickland et al., 2009; Blagodatsky et al., 2010; Garcia-Pausas and Paterson, 2011), and could potentially explain the large variation in PEs among different studies. Differences in environmental factors (e.g. soil moisture, pH) and SOM quality (e.g. chemical recalcitrance and C:N ratio) are also of importance for SOM decomposition rates, meaning variations in one or several of these biotic and abiotic factors are likely to have a strong effect on the magnitude and direction of PEs. Finally, plant species differences in quality and quantity of root exudates are thought to be key plant traits controlling the PEs (Wang et al., 2016b). In this section, I summarized the current knowledge about how some of these factors influence PE.

Microbial activity, biomass and community composition

Microbial activity and biomass The first step in the microbial utilization of SOM is its decomposition into compounds that can be taken up and metabolized by microorganisms. Variations in the biomass and activity of microorganisms will, therefore, result in corresponding variations in the turnover rate of SOM (Blagodatsky et al., 2010; Thiessen et al., 2013). In other words, a change in microbial activity and/or biomass, caused by the input of labile C that can serve as the primary energy source, can potentially be responsible for the PE (Kuzyakov, 2002; Cheng and Kuzyakov, 2005). Accordingly, many studies have found a significant relationship between PEs and microbial activity and biomass (Xiao et al., 2015; Wang et al., 2016a; Fang et al., 2018; Li et al., 2018), and a review by Kuzyakov et al. (2010) concluded that the most important mechanism for regulating PEs is the acceleration or retardation of SOM turnover due to increased activity or amount of microbial biomass. This effect might be linked to increased microbial enzyme production since enzyme production is strongly correlated to the microbial biomass and activity (Albiach et al., 2000).

Microbial community composition Variations in the PEs might be caused by changes/variations in the microbial community composition (Blagodatskaya and Kuzyakov, 2008; Su et al., 2017). It is commonly proposed that PE involves a succession of processes that are partly connected to the succession of microbial communities and their functions (Blagodatskaya and Kuzyakov, 2008). The addition of labile C may first stimulate

18

the growth and metabolism of microbial r-strategists, followed by a gradual increase in the abundance of K-strategists (Fontaine et al., 2003). Bacteria are generally considered to be r-strategists, while fungi are considered as K-strategists. According to this theory, the initial microbial community response to the addition of easily available substrates is the activation of bacterial metabolism, which results in apparent PE, while fungal activity increases over time and is responsible for the later phase when real PEs occur (Blagodatskaya and Kuzyakov, 2008). In addition, fungi can grow better than bacteria in low nutrient zones since they can access and degrade substrates that are not available to most bacteria, and acquire nutrients from distantly-located substrates by their hyphae (Chalot and Brun, 1998; Otten et al., 2001). Thus, the dynamic of long-term PEs can possibly be explained by successional changes in microbial community composition, manifested as a change in the fungal/bacterial ratio (Fontaine et al., 2003). Accordingly, several studies have demonstrated that the PE is accompanied by such changes in the composition of the soil microbial community (Fontaine et al., 2011; Bastida et al., 2013; Zhao et al., 2018).

Plant species and physiology Processes in the rhizosphere are primarily controlled by interactions between plant roots and the microbiological community (Cheng et al., 2014). Plant roots influence the soil's physical, hydrological, and chemical environment (e.g. SOM content, water content, pH, etc.) and associated biological processes, all of which are important for the turnover of C and other nutrients (Gregory, 2006). Part of the plants recently photosynthesized C is allocated belowground to root exudates, thereby influencing the microbial metabolism. Therefore, the SOM decomposition rate is strongly influenced by plants and their rhizodeposits (Cheng and Kuzyakov, 2005). In the following section, I summarized the current knowledge about how plant species, plant phenology and plant age influence PEs.

Plant species Many studies have shown that decomposition and nutrient release from SOM varies among plant species, and different plant species induce different rhizosphere PEs (Zhu et al., 2014; Wang et al., 2016b; Xu et al., 2017; Yin et al., 2018). For example, soybean induced higher PEs than that of wheat when the two species were grown under the same environmental conditions (Cheng et al., 2013). A meta-analysis found that woody species induce the highest PEs followed by grasses, while crops induce the lowest level of PEs (Huo et al., 2017). The response of the PE to environmental changes also seems to differ among plant species. For instance, Xu et al. (2017) found that elevated CO2 increased the PE induced by white lupin by 47%, but decreased the PEs induced by wheat by 22%. They suggested that changes

19

in the quality and quantity of root exudate in response to elevated CO2 might account for the species-specific responses in PEs.

In addition, a study with different tree species showed that up to 60% of plant photosynthates C are transferred to the soil by mycorrhizal hyphae (Godbold et al., 2006). Mycorrhizal partners, which vary in their associations with different plant species (Phillips and Fahey, 2006), might, therefore, be an important factor that influences the PE

Plant phenology and plant age Several studies have found that the magnitude of the PE differs among different plant phenological stages (Fu and Cheng, 2002; Cheng et al., 2003). For example, a study with spring wheat found that the PE of wheat reached 287% at the flowering stage and declined significantly afterward (Cheng et al., 2003). Furthermore, although young roots exude organic substances more intensively than old roots (Warembourg and Estelrich, 2001), the fully developed plant roots can occupy more soil space than the younger ones. This might be the reason why older plants cause stronger PEs, at lease when expressed on soil weight or soil volume basis. In line with this proposal, Kuzyakov et al. (2001) found PEs induced by Lolium to increase during a 3.5 months cultivation period. In the first half of the observation period, a small non-significant negative PEs was observed, but the PEs became positive after 2.5 months of vegetative growth. Other studies have found similar results.

Soil properties As mentioned above, SOM decomposition and PE are mediated by microorganisms, whose activity, in turn, is influenced by various soil properties.

In this section, I reviewed the current knowledge about how some of the most influential soil properties influence microbial SOM decomposition and PEs.

Soil type SOM is made up of a heterogeneous mixture of compounds of varying chemical complexity that contain nutrients essential for microorganisms and for plant growth (Lal, 2009; Schmidt et al., 2011). Accordingly, the rhizosphere PE has been found to differ between different soil types. For example, soybean and sunflower caused a positive PEs when grown in agricultural soil, but no PEs when grown in annual grassland soil under the same environmental conditions (Dijkstra et al., 2006). The authors suggested that rhizosphere PE of SOM decomposition is more likely to occur in soils of relatively high fertility that can sustain high plant productivity, and possibly in soils where the native SOM contains less labile C compounds. Zhang et

20

al. (2017) found distinctly different PEs in a Millisol and an Alfisol. Positive PEs occurred in the Mollisol, while negative PE was found in the Alfisol. They suggested that the differences in the PE between two soils largely stem from differences in the soil clay content.

Soil C is generally the driving force of most microbially mediated processes (Demoling et al., 2007). Accordingly, a laboratory experiment showed that a fine loamy Gleyic Cambisol containing 4.7 % organic C induced 30- to 100-times higher PE when compared to a sandy Gleyic Camisole with 0.7 % organic C (Kuzyakov et al., 2001). In contrast, the relative PE was 1.1-3.3 times stronger in mineral soil than in the organic layer in a study by Wang et al. (2015). These contrasting results demonstrate that the importance of organic C for regulating the PEs not straight forward, and that other soil and/or plant properties are at least equally important to consider. For example, grasses allocate a larger portion of net assimilated C belowground and exude more C to the rhizosphere in soils poor in organic C, than in soils rich in organic C (Warembourg and Estelrich, 2001). A probable reason is that plants stimulate microorganisms to degrade SOM and mobilize nutrients by increasing their root exudation rate in poor soils. Therefore, plants and microorganisms in soils that are rich in organic C are less dependent on increasing SOM decomposition to obtain nutrients.

Soil pH Soil pH can affect almost all chemical, physical and biological soil properties (Kemmitt et al., 2006), and is often regarded as one of the most important factors regulating the microbial biomass, activity and community composition (Bååth and Anderson, 2003; Aciego Pietri and Brookes, 2009). Soil pH has a particularly strong influence on the relative abundance and growth of fungi and bacteria (Rousk et al., 2009), meaning that soil pH could also potentially be an important regulator of the PE. Accordingly, an analysis of six previously published studies revealed a linear increase in the priming efficiency (the PE expressed as a percent of added C) in the pH range 2.7 to 7.8 (Blagodatskaya and Kuzyakov, 2008). Likewise, Aye et al. (2017) observed that the PE caused by the addition of plant residues increased as a result of long-term liming of acid soils. However, another recent study by Aye et al. (2018) showed that PE was not related to soil pH during a 30-day incubation period. In this study, the highest PE occurred at pH 4.1, the second-highest at pH 6.6, and the lowest at pH 4.7. The authors suggested that the high PE in pH 4.1 soil could be ascribed to the greater net increase in microbial biomass carbon in response to labile C input, which was because the net increase of microbial biomass carbon in response to C-substrate treatment69% at pH 4.1 was, 29% at pH 4.7 and 48% at pH 6.6.

21

Soil nitrogen Nitrogen is the most commonly limiting nutrient of primary production and is thus of crucial importance for regulating the C balance of terrestrial ecosystems. Nitrogen also influences belowground processes and PE. As described above, positive PE may occur if the soil is rich in available C but low in available N, due to microbial N mining (Kuzyakov et al., 2000). On the other hand, if N is abundant this can lead to negative PEs due to preferential substrate utilization (Cheng, 1999). At severely N limiting conditions, competition for N between microorganisms and plants might also result in negative PEs (Kuzyakov, 2002). However, the effect of N fertilization on the PE is contentious, and different studies have found decreased (He et al., 2016; Aye et al., 2018), enhanced (Chen et al., 2014; Lu et al., 2018) or no effect (Qiao et al., 2016; Di Lonardo et al., 2019) of N fertilization on the PE. These discrepant results might be caused by variations in C substrate quality, the amount of N applied relative to the indigenous soil N content, and other physical, chemical and biological soil properties. Hence, the role of N in regulating the PE still warrants further investigations.

Differences in the soil C:N ratio might be one factor that could explain why N fertilization has such contrasting effects on the PE. Kuzyakov (2002) have proposed four scenarios that determine how the SOM decomposition rate can change depending on the amount of decomposable organic C and N in soil: (1) The competition for N between plants and microorganisms will become dominant when both C and N are simultaneously limited in the soil. This would inhibit microbial activity and biomass, leading to decreased SOM decomposition. (2) If available C is limiting microbial activities, this would result in preferential substrate utilization where microorganisms preferentially use labile C (rhizodeposits) instead of decomposing SOM to gain C, resulting in decreased SOM decomposition. (3) In cases where labile C is added to soils with low C:N ratio, microorganisms are activated by the input of labile C and use it to decompose SOM for acquiring additional N (4) When both available C and N are sufficient in the soil, no changes in SOM decomposition rates will be observed, since microorganisms have the optimal nutritional conditions and do not need to acquire additional nutrients from SOM when labile C is added.

Soil water content Soil moisture can significantly influence the PE through its impact on microbial activity, plant growth, and root exudation rates (Dijkstra and Cheng, 2007; Dijkstra et al., 2010). Root exudates may be more effective in stimulating microbial activity and decomposition in wet than in dry soils, because of more rapid diffusion of exudates away from the root and reduced chances of exudates being actively re-adsorbed by roots (Jones and Darrah, 1993). For example, Dijkstra et al. (2010) suggested that when water availability is low, plant species complementarity and

22

selection effects on water and N use can decrease soil C decomposition and the rhizosphere PE, based on results from a greenhouse experiment with five semi-arid grassland species and two water levels. In contrast, Lu et al. (2018) found that high soil moisture reduced the rhizosphere PEs. They suggested that this was caused by increased microbial nutrient limitation in the high soil moisture treatment. High moisture increased the plant biomass and plant N uptake, which might have resulted in the plant-microbe competition for mineral N and decreased the microbial activity, and hence depressed SOM decomposition. It is not clear why this would be the case, but the contrasting results demonstrate that there are no straightforward relations between water content and PEs

Amount of the ‘primer’ The amount of the “primer”, i.e. the labile C that induces the PE, is considered an important regulator of PEs. Higher amounts of labile C input will result in a more pronounced stimulation of microbial activity, resulting in stronger PEs (Guenet et al., 2010; Bengtson et al., 2012). Therefore, the rate of available C input to the soil will regulate the magnitude of the PE. However, it appears as if there is an upper limit for this effect. For example, Bengtson et al. (2012) found that each mg of exuded C resulted in decomposition and release of 6 mg bioavailable C from SOM, but there was no or little additional increase in the relative PE when the root exudation rates exceeded 16 mg C per kg soil per day. This suggests that other factors also need to be taken into consideration. For example, Blagodatskaya et al. (2008) suggested that the amount of added substrate relative to the amount of microbial C is a key factor regulating PEs, rather than the amount of substrate on its own. Furthermore, a combined addition of different substrates can induce a higher positive PE than single substrate additions (Hamer and Marschner, 2005). The authors further suggested that the substrate concentrations, as well as the adaptation of microorganisms to the added substrate, are important factors controlling PEs.

23

Aim and approaches of studies

This section of my thesis is focussed on describing the aims and approaches of Paper I-V, while the methodology is described in greater detail in the next section.

Paper I Most experiments aimed at investigating the rhizosphere PE in the presence of living roots have been performed using unplanted soil as the control, which could potentially complicate quantification of the rhizosphere PE and mislead our understanding of the underlying mechanisms. In order to evaluate the best choice of control to achieve the aim of this thesis, I performed an experiment in with Scots pine (Pinus sylvestris) and Norway spruce (Picea abies) seedlings, and two different controls, 1. a never planted control (unplanted control), and 2. A control where planted seedlings were cut at the soil surface five days before the measurement (planted control). The aim was to determine how the choice of control will influence the calculated rhizosphere PEs, as well as to test if the variation of the calculated rhizosphere PE can be linked to biological and environmental differences between the two controls.

The rhizosphere PE of SOM decomposition and N mineralization, microbial growth rate, and community composition, as well as the potential enzyme activity, were measured to achieve the aim of the study. The 15N pool-dilution technique was used to determine the gross N assimilation and mineralization rate (Davidson et al., 1991), and the rhizosphere PEs of SOM decomposition and gross N mineralization were calculated as in Bengtson et al. (2012). Microbial community composition was determined by phospholipid fatty acid (PLFA) analysis (Frostegård and Bååth, 1996), and the fungal and bacterial growth rate was determined using isotope incorporation methods (Bååth, 1994; Bååth, 2001). The potential enzyme activity was determined by using a combination of fluorometric and photometric methods (Marx et al., 2001; Saiya-Cork et al., 2002).

24

Fig. 2. A) 13CO2 labelling and the Picarro analyze used for measurements 13C/12CCO2 measurements. B) Pine and spruce seedlings cultivated in the greenhouse and then used in the experiments.

Paper II The objective in Paper II was to investigate if the rhizosphere PE differs between two tree species (Norway spruce and Scots pine) and if their rhizosphere PE responds differently to different light intensities. Plant photosynthetic intensity, which is a main controlling factor of belowground plant C allocation and root exudation rate (Kuzyakov and Cheng, 2001), is a possible reason why the rhizosphere PE varies widely among plant species. I, therefore, hypothesized that the rhizosphere PEs would differ between the two tree species under the same environmental conditions since they have different light saturation and compensation points and different rhizosphere environments. I further hypothesized that light intensity would influence the rhizosphere PEs through its effect on the photosynthetic intensity and root exudation rates. The hypothesis was tested in a greenhouse experiment with pine and spruce seedlings that were cultivated for two months before the measurements.

The rhizosphere PEs of SOM decomposition and gross N mineralization was determined as described in the paper I above. A 13CO2 pulse-chase method was used to determine the root exudation rate, as in Bengtson et al. (2012). I also measured the microbial growth rate and microbial community composition as in paper I, to investigate their role in rhizosphere PEs.

25

Fig. 3. Pots with the two tree species (Norway spruce and Scots pine) used in the experiment, with the different number of intact seedlings remaining.

Paper III In my third experiment, I cultivated the same tree species (Norway spruce and Scots pine) as in Paper II, but this time for 6 months. Five days prior to measurement of the root exudation rate and the PE, some of the seedlings were cut at the surface of the soil, leaving 0, 1, 3 or 5 intact seedlings in each pot (Fig. 3). The amount of the primer is considered as an important regulator of PEs (Bengtson et al., 2012), and I hypothesized that higher root exudation rates in pots with higher numbers of intact seedlings would result in stronger stimulation of microbial SOM decomposition and gross N mineralization rates, and consequently a stronger rhizosphere PE. I further hypothesized that the rhizosphere PE is caused by increased microbial enzyme production and that this would be reflected in higher potential enzyme activity in treatments with high rhizosphere PEs. A stable isotope probing (SIP) experiment measuring the incorporation of deuterium (D) (derived from D2O) and 13C (derived from root exudates) into microbial biomarker PLFAs was also performed to identify which microbial groups that grew on root exudates and/or on resources released as a result of the rhizosphere PE. SOM decomposition, gross N mineralization, root exudation rate, fungal and bacterial growth rates, and the rhizosphere PE was determined in the same way as in Paper I and II.

26

Fig.4. Spruce seedlings cultivated under controlled CO2 concentrations.

Paper IV Belowground C allocation by plants can be affected by the CO2 concentration (Ainsworth and Long, 2005), which might thus be an important regulator of the PE. As stated above, soil N availability has also been proposed as a key role in regulating the PEs (Dijkstra et al., 2013), and it is possible that plant-mediated belowground responses to elevated CO2 are dependent on the amount of plant available N (Kuzyakov and Blagodatskaya, 2015). In this experiment, I, therefore, quantified the combined effect of N fertilization and elevated CO2 on the root exudation rate and rhizosphere PEs. I hypothesized that elevated CO2 would increase the root exudation rate and result in higher rhizosphere PEs. This hypothesis was tested by exposing spruce seedlings to ambient (400 ppm) and elevated CO2 (600 ppm) environment for 7 days (Fig. 4). The CO2 treatments were combined with three levels of N fertilization, to test if any changes in the root exudation rate and rhizosphere PEs at elevated CO2 are moderated by the N availability. The relationship between rhizosphere PEs and plant N uptake was also tested. SOM decomposition, gross N mineralization, root exudation rate and the rhizosphere PEs were determined in the same way as above.

27

Paper V Plant available N and microbial available N is produced by the breakdown of soil polymers (e.g. chitin and protein) that represent a major fraction of soil N. Soil microorganisms may adjust the production of enzymes that target different polymers, resulting in a balanced supply of available C and N (Burns et al., 2013; Bell et al., 2014). In this experiment, I first hypothesized labile C addition would increase microbial N demand and hence stimulate the depolymerization of lignin, chitin, and protein. Although lignin does not contain N, lignin depolymerization can still increase soil N availability (Carreiro et al., 2000; Schimel and Bennett, 2004; Knicker, 2011). Secondly, I hypothesized that the opposite effect would occur after the addition of available N.

The hypotheses were tested by using a combination of two complementary ten-day incubation experiments. In experiment 1, the soil was amended with 13C enriched glucose and available N (ammonium). In experiment 2, the same soil was mixed with lignin, chitin, or protein before the addition of available C or N as above. In the second experiment, glucose had natural 13C content, while lignin, chitin, and protein were enriched 13C. Microbial community composition and potential enzyme activity were also measured to the hypotheses.

29

Methodology

This chapter provides a short overview of the methods used to achieve the aims and test the hypotheses in paper I-V.

Priming of N mineralization In paper I, II, III and IV, gross N mineralization was determined using the 15N-pool dilution method (Davidson et al., 1991). Briefly, 15N enriched NH4Cl was injected into the soil at different positions. Half of the replicates were harvested within 2 hours after the 15NH4Cl addition (T0) and the rest 24 hours after the 15NH4Cl addition (T1). The concentration of 15NH4

+-N and 14NH4+-N at T0 and T1 was

determined by isotope ratio mass spectroscopy (IRMS) after extraction and diffusion according to standard procedures (IAEA, 2001). Then the gross N mineralization and assimilation rates were calculated using the FLAUZ model (Mary et al., 1998), from the differences in the concentration of N and 15N of ammonium between T0 and T1.

The PE of N mineralization was calculated as:

Nprimed = N mineralizationtreatment–N mineralizationcontrol

Nprimed (%) = – × 100 (1)

Where N mineralization treatment is the gross N mineralization in treatment samples, and N mineralization control is the gross N mineralization in control samples.

30

Priming of SOM decomposition In Paper I, II, III and IV, SOM decomposition was calculated from N assimilation measured as above, using the following equations:

Cassimilated = Nassimilated × f 15Nmic × C: Nmicroorganisms (2)

Where Cassimilated is the microbial C assimilation, Nassimilated is the total N assimilation by plants and microorganisms calculated from the 15N-pool dilution method above, f 15Nmic is the fraction of the total assimilated 15N that was assimilated by the microorganism, and C:Nmicroorganisms is the soil microorganism C: N ratio.

SOMdecomposed= (3)

Where SOMdecomposed is the SOM decomposition and CUE is microbial C-use efficiency.

Then the PE of SOM decomposition was calculated:

SOMprimed = SOM decompositiontreatment – SOM decompositioncontrol

SOM primed (%) = – ×100 (4)

Where SOM decompositionreatment is the SOM decomposition in the treatment with intact plants and SOM decompositioncontrol is the SOM decomposition in the control treatment with cut plants.

In order to correctly represent the uncertainty of the assumed variables (C:Nmicroorganisms and CUE), I performed the calculations in @RISK 7.6 (Palisade Corporation, Ithaca, NY, USA) using a probabilistic Monte Carlo analysis. Significant differences among treatments were assessed by testing if the 85% confidence intervals of these calculations overlapped (Payton et al., 2000; Payton et al., 2003).

In Paper V, the respiration and isotopic composition of respired CO2 were measured immediately and then during 10 days after the start of the experiment. Briefly, gas-tight flasks equipped with a septum was evacuated four times and refilled with pressurized air of known concentration and isotopic composition of CO2. Empty flasks were processed in the same way and used as blanks. The flasks were incubated for between 4-8 hours before analysis of the concentration and isotopic composition of accumulated CO2. The analyses were performed using a Delta Ray Isotope Ratio Infrared Spectrometer (Thermo scientific), calibrated against two CO2

31

standards of different isotopic composition. The CO2 concentration and isotopic composition were corrected for the initially present CO2 using the values derived from the blanks. We then distinguished CO2 from different sources using the following equations: 𝐶 = 𝐶 × % %% % (5) 𝐶 = 𝐶 − 𝐶 (6)

Where CTotal was the concentrations of total CO2, C1 and C2 were the concentrations of CO2 from source 1 and from source 2, respectively. The atom%Total, atom%1 and atom%2 were the corresponding 13C contents (in atom% 13C) of respired CO2.

In experiment 1, source 1 was the added 13C enriched glucose and source 2 was the natural abundance of 13C in soil organic C. In Experiment 2, source 1 was the added 13C enriched polymers (lignin, chitin, or protein), while source 2 included both natural abundance of 13C in soil organic C and the added glucose.

The concentration and 13C content of microbial C was measured in 0.5 M K2SO4 extracts (after chloroform fumigation) and 13C in dissolved organic C in 0.5 M K2SO4 extracts of fresh soil, using an LC Isolink IRMS system. To assess the effect of C and N addition on the mobilization of C from SOM in experiment 1, and from polymers (lignin, chitin and protein) in experiment 2, the sum of SOM-derived C and polymer-derived C in dissolved organic C, microbial C and cumulative respiration from SOM and polymers, was calculated (as shown in Fig. 1 in paper V).

Root exudation rate In Paper II, III and IV a 13CO2 pulse-chase method was used to estimate the root exudation rate, as in Bengtson et al. (2012). Seedlings were labelled with 13CO2 by placing them in a plexiglass chamber (Fig.2A). 10 mL (in paper II) or 20 mL (in paper III and IV) 13CO2 (99 atom % 13C) was injected into the chambers and the concentrations of 12CO2 and 13CO2 were measured during the next 1.85 hour, using a Picarro G2201-i analyzer (Picarro Inc., Santa Clara, CA, USA). The seedlings were then destructively harvested five days after the 13CO2 labelling and soil samples were freeze-dried for analysis of the soil 13C content in a Flash 2000 elemental analyzer (Thermo Scientific Inc., Bremen Germany).

The root exudation rate was calculated based on the soil 13C content:

32

Root exudation rate = × ( / ) / ( ) / (9)

Where 13Csoil is the amount of CO2 derived 13C that recovered in soil, f 13C lost is the fraction of 13C lost from the soil due to microbial respiration, t1 is the duration of the light period (12 hours), t2 the length of 13CO2 labelling (1.85 hours), and 13C-CO2 the atom percent of 13C excess in CO2. The short duration of the 13CO2 pulse-labelling was to ensure that there was no significant dilution of the 13CO2 label caused by plant or microbial respiration and that no recycling of fixed 13C that occurred during the pulse-labelling time. The calculation in equation 9 was performed in @RISK 7.6 as mentioned above. I used 0.71±0.09 as the value of the 1- f 13C lost, based on Bengtson et al. (2012).

Microbial growth rate I measured bacterial and fungal growth rates in Paper I, II, III and IV, to investigate how the fungal and bacterial activity influences the PEs

The bacterial growth rate was estimated by the 3H-leucine incorporation method. The method is based on the fact that bacteria will take up and incorporate amino acids (e.g. leucine) into their proteins at a rate corresponding to the growth rate. Therefore, the bacterial growth rate can be indicated by the amount of incorporated leucine before and after incubation with isotopically labelled leucine (Bååth, 1994). Briefly, a diluted 3H-leucine solution is added after the extraction of bacteria from soil. Washing and subsequent measurement of radioactivity are performed after incubation for 2 hours, and the amounts of leucine derived 3H measured using a liquid scintillation counter (Perkin-Elmer Life Sciences). The amount of 3H-leucine incorporated into bacteria per hour per gram of soil was then used as a measurement of bacterial growth, after conversion to bacterial biomass C production.

The fungal growth rate was estimated by the 14C-acetate incorporation method, as in Bååth (2001). Quantifying fungal biomass can be performed by estimating a substance present in fungal structures. Ergosterol is a membrane lipid (a sterol), which does not occur in considerable amounts except in fungi. Therefore, the rate of synthesis of ergosterol can be used as an indicator of fungal growth rate. In order to estimate ergosterol synthesis, I used 14C-acetate, which is a precursor of lipids. Briefly, fresh soil was transferred to test tubes a 14C-acetate solution added. Formalin was added to terminate the fungal growth after incubation for 4 hours. Ergosterol was then extracted, separated, and the amount of incorporated radioactivity determined. The amount of 14C-acetate incorporated into fungal

33

ergosterol per hour per gram of soil was used as a measurement of fungal growth, after conversion to fungal biomass C production.

Microbial community composition In Paper I, II, III, IV and V, microbial community composition was determined by measuring microbial biomarker phospholipid fatty acids (PLFAs) (Frostegård and Bååth, 1996; Bååth and Anderson, 2003). Phospholipids occur in all living microorganisms and degrade fast when the organism dies. The total amount of microbial phospholipids can, therefore, be used as a measure of the living microbial biomass (Frostegård and Bååth, 1996). The fatty acid part in phospholipids differs among microorganisms, and fatty acids can, therefore, be used to distinguish different groups of microorganisms, like fungi and bacteria. I also used stable isotope (13C) probing (SIP) to investigate the incorporation of isotopes into different microbial biomarker PLFAs. In paper III, a combination of 13C and D SIP was performed, which enabled me to determine which microbial group that was growing directly on the root exudates, and on resources released from SOM by the PEs. I assumed that microorganisms that assimilated D, but not 13C, was responsible for the observed PEs (Hungate et al., 2015). Microbial incorporation of 13C and D into PLFAs were calculated from the atom% 13C and atom% D excess in individual PLFAs.

Potential enzyme activities In Paper I, III and V, I investigated if increased enzyme production and/or activity is the mechanism underlying PEs. The potential activity of six extracellular hydrolytic enzymes and two oxidative enzymes (Summarized in Table 1) were measured by using a combination of fluorometric and photometric methods (Marx et al., 2001; Saiya-Cork et al., 2002).

Fresh soil was amended with fluorescence-labeled substrates to analyze hydrolytic enzyme activities (Table 1). Fluorescence was then measured after incubation to determine the concentrations of released 4-methylumbelliferone (MUF) or 7-amido-4-methylcoumarine (AMC), respectively. Potential hydrolytic enzyme activities were expressed in nmol MUF or AMC per gram dry soil per hour.

Potential phenoloxidase and peroxidase enzyme activity were mixed with L-3,4-dihydroxyphenylalanin as the substrate. Samples for measurements of peroxidase activity also receive 10 µl 0.3% H2O2 (Table 1). The absorbance was then measured

34

before and after 20 hours incubation. Potential oxidative enzyme activity is expressed in nmol DOPA per gram dry soil per hour.

Table 1. Extracellular enzymes and their abbreviations, substrate, and their function for the degradation of macromolecules.

Enzyme Substrate Degradation of macromolecules α-1,4-glucosidase 4-MUB-α-D- glucosidase Maltose β-1,4-xylosidase 4-MUB-β-D- xylosidase Hemicellulose β-1,4-glucosidase 4-MUB-β-D- glucosidase Cellulose and cellobiose L-leucine aminopeptidase L-leucine-7-amino-4-

methylcoumarin Protein and peptides

β-1,4-N-acetylglucosaminidase

4-MUB-N-acetyl-β-D-glucosaminide

Chitin and bacterial peptidoglycan

Phenoloxidase L-3,4-dihydroxyphenylalanine Phenolic compounds Peroxidase L-3,4-dihydroxyphenylalanine and

H2O2 Peroxides (e.g.Hydro peroxides and lipid peroxides).

35

Main findings

Evaluation of different controls In paper I, I compared two kinds of controls (unplanted and planted) to investigate how the choice of control influences the calculated rhizosphere PE. The unplanted control consisted of soil that had never been planted and the planted control of pots with soil where the seedlings had been cut at the soil surface five days before the measurements.

My results showed that the rhizosphere PE varied widely depending on the control. In the pine treatment, the rhizosphere PE of SOM decomposition did not differ significantly when calculated using the planted control and unplanted control, respectively, while rhizosphere PE of gross N mineralization was much higher when calculated from the unplanted control. In the spruce treatment, the rhizosphere PE of SOM decomposition was negative when calculated using the planted control, and positive when calculated from the unplanted control. As in the pine treatment, rhizosphere PE of gross N mineralization was much higher when calculated from the unplanted control. (Fig. 1 in paper I). The variation in the rhizosphere PE was caused by lower SOM decomposition and gross N mineralization rates in unplanted than planted control. The reduction seemed to be caused by lower microbial biomass, lower bacterial and fungal growth, and higher bacterial dominance in unplanted control, but possibly also by lower inorganic N concentrations in the planted control compared to the unplanted control (Fig. 1 and Table 1 in paper I).

I also performed an experiment with 13C labelled plants, which showed that a 5-day time frame that cut seedlings above-ground parts was sufficient to let root activity cease and short enough to prevent any PEs induced by decomposed dead roots (Fig. 2 in paper I). Based on these experiments, I found that the planted control, where above-ground plant parts removed only a few days before measurement the rhizosphere PE, was a better choice for achieving the aims of my thesis and investigate the direct effect of living roots on belowground SOM decomposition.

36

Plant species effects on rhizosphere PEs In paper II, my results demonstrated that spruce induced positive rhizosphere PEs, while pine induced a negative rhizosphere PEs (Fig. 2 and 4 in paper II). I suggested that the negative rhizosphere PE in the pine treatment was caused by the competition for available N between plant and microorganisms (Kuzyakov and Xu, 2013). Accordingly, we found that both the inorganic N concentration and the gross N mineralization rate were lower in soil with pine seedlings than in soil with spruce seedlings (Table 2 and Fig. 4 in paper II).

I concluded that higher N availability in combination with input of labile C by root exudates might have increase microbial decomposition of SOM as well as gross N mineralization rates in the spruce treatment, while available N was not enough for this occur in the pine treatment, resulting in a competition for N between plants and microorganisms and a negative rhizosphere PE. Accordingly, my results showed that there was a significant positive relationship between inorganic N concentration and rhizosphere PEs (Fig. 3 in paper II).

In paper III the presence of living plants did not influence the gross N mineralization rate, meaning that no rhizosphere PE of N mineralization occurred (Table 2 in paper III). Possible reasons include more efficient microbial use of already available N (Wild et al., 2017), due to decreasing inorganic N concentrations as the experiment progressed (Table 1 in paper III). Accordingly, more efficient microbial recycling of N occurs in response to low N availability (Bengtson and Bengtsson, 2005).

Even if no rhizosphere PE of gross N mineralization occurred in paper III, pine seedlings induced a positive rhizosphere PE of SOM decomposition while spruce seedlings induced a negative rhizosphere PE of SOM decomposition (Table 2 in paper III). My results suggested that the negative rhizosphere PE that occurred in the spruce treatment was caused by an opportunistic subset of the microbial community that grew on root-derived labile C compounds, while at the same reducing the activity of SOM decomposers by depleting available N. For some reason, this did not appear to occur in the pine treatment, where available N was sufficient for root exudate to stimulate fungal decomposers and induce positive rhizosphere PEs, despite low inorganic N concentrations. I discuss the possible reasons below.

Interestingly, my results in paper II and III showed opposite rhizosphere PEs of the two tree species. The contradictory findings might have been caused by changes in soil N availability and microbial community composition during the cultivation of the seedlings. The seedlings were cultivated for two months in paper II and for six months in paper III, and as previously discussed plant age is a possible factor regulating PEs (Cheng et al., 2003). Furthermore, due to the longer cultivation time

37

in paper III, the soil inorganic N content decreased from 1.19 µg N g-1 dw soil in paper II, to 0.47 µg N g-1 dw soil in paper III in the spruce treatment, (Fig. 5A). This decrease might have induced competition for N in paper III, resulting in suppressed activity of microbial decomposers and thus negative PEs. Accordingly, I found that addition of N fertilizer changed the rhizosphere PEs from negative to positive in paper IV (Table 1 in paper IV), where the seedlings had been cultivated for 4 months. It is likely that the N deficiency was even more pronounced after 6 months of cultivation, as in paper III.

The soil inorganic N content decreased also in the pine treatment, but not to the same low level as in the spruce treatment (from 0.98 to 0.68 µg N g-1 dw soil). For some reason, the decrease in inorganic N resulted in the opposite effect compared to the spruce treatment. My analyses of the microbial community composition might provide a clue to why this was the case. There was a much stronger fungal dominance in the pine treatment in paper III than in paper II (Fig. 5B). This microbial community shift from bacterial dominance to fungal dominance occurred due to the difference in cultivation time between the two experiments. Fungi have higher C:N ratios than bacteria and are better adapted to low N conditions (Sterner and Elser, 2002). The shift in microbial community composition may, therefore, have relieved the N limitation experienced by the microbial decomposer community in paper II. The increased abundance of fungal biomarker PLFAs might also be a result of increased ectomycorrhizal colonization of the pine seedlings as time progressed since it has been suggested that ectomycorrhiza might be an important player in the regulation of rhizosphere PEs (Zak et al., 2019). However, my data does not allow me to evaluate if this was the case. Finally, pine is better adapted to low nutrient concentrations than spruce (Merilä and Derome, 2008), and it can be speculated that pine roots selected for a microbial decomposer community that is equally well adapted to low nutrient concentrations.

38

Fig. 5. (A) The relationship between priming of SOM decomposition and soil inorganic N content, (B) Principal component analysis (PCA) of relative individual PLFAs into total microbial biomarker PLFAs in all treatments in paper II and III.

In conclusion, paper II and III showed that the magnitude and direction of the rhizosphere PE depend on plant species and is also affected by the interaction between N availability, plant roots, and microbial community composition.

Light intensity effect on rhizosphere PEs As stated above, pine induced no or negative rhizosphere PEs on both SOM decomposition and gross N mineralization in paper II. The negative rhizosphere PE of SOM decomposition tended to be more pronounced and was only significantly different from zero under low light (Fig. 2 in paper II). The seedlings also allocated more C belowground to root exudates in the low light treatment, which might have induced stronger competition for N between microbial decomposers and plant roots and their associated ectomycorrhiza. This is consistent with my observation that the soil inorganic N increased with 19% after cutting the pine seedlings in the low light treatment, while it only increased with 12% in the high light treatment (Table 2 in paper II). Likewise, cutting the seedlings resulted in a more pronounced increase in microbial biomass in the low light than in the high light treatment, as indicated by the change in total abundance of microbial PLFAs (Table 3 in paper II).

In the spruce treatment, the positive rhizosphere PEs on SOM decomposition and gross N mineralization were both more pronounced in the low than in the high light treatment (Fig. 2 and 4 in paper II). This was possibly caused by the higher

39

concentration of available N in low light treatment. Sufficient inorganic N to sustain enzyme production, possibly in combination with root exudates that stimulate the extracellular oxidative enzyme activity (Bengtson et al., 2012; Rousk et al., 2015), might be the reason.

In brief, my results showed that the light intensity is an important regulator of the rhizosphere PEs, but also that its effect is plant species dependent.

Elevated CO2 and N fertilization effects on rhizosphere PEs In paper IV, rhizosphere PEs on SOM decomposition was positive in the high N fertilization treatment, and negative in the no N and low N fertilization treatments, irrespectively of the CO2 concentration (Table 1 in paper IV). In the elevated CO2

treatment, there was no or very low rhizosphere PE on gross N mineralization, regardless of the N fertilization level, while N fertilization resulted in positive rhizosphere PE on N mineralization in the ambient CO2 treatment (Table 2 in paper IV). The results could not be linked to an effect of elevated CO2 and N fertilization on the root exudation rate. The findings instead suggest that the effect of N-fertilization and elevated CO2 on the partitioning of N between plants and microorganisms are more important for SOM decomposition and rhizosphere PEs than the root exudation rate itself. The results also showed that N-fertilization in combination with elevated CO2 stimulated microbial decomposition of SOM to a greater extent than photosynthesis, meaning that N fertilization in combination with elevated CO2 can shift the C balance and result in a reduction of soil C stocks in N limited forest ecosystems.

Root exudate effects on the rhizosphere PE In paper II, my results showed that the light intensity can regulate the root exudation rate, depending on the plant species. Pine seedlings kept under low light intensity exuded more C rate than those kept at a high light intensity, while no difference was found in the root exudation rate of spruce seedlings kept under low and high light intensity (Fig. 1 in paper II). I can only speculate why pine seedlings allocated more C to belowground at low light intensity, but a possible reason is that shaded pine seedlings will invest more in belowground symbionts, so that it can benefit from the transfer of water, C, and N from larger trees through a common mycorrhizal network (Simard et al., 1997; Beiler et al., 2010; Teste et al., 2010). This did not occur in spruce, possibly since spruce achieves optimal photosynthesis

40

in a lower range of light intensities than pine (Kurets et al., 1994). Regardless, the variation in root exudation rate among treatments could not explain the different rhizosphere PEs in this experiment. As discussed above, it instead appears as if variations in soil N availability was the reason for different rhizosphere PEs between the two species and light intensities.

In paper III, I estimated the root exudation rate in pots with a varying number of pine and spruce seedlings. My results showed that the root exudation rate increased with the number of intact seedlings. It was 213% higher in pots with five seedlings compared to pots with one seedling in the pine treatment, and 108% higher in pots with five seedlings compared to pots with one seedling in the spruce treatment (Fig.1 in paper III). However, higher root exudation rates did not result in a corresponding increase in the rhizosphere PEs. Conversely, the rhizosphere PE of SOM decomposition was lower in the pine treatment with 5 intact seedlings compared to the treatments with 1 or 3 intact seedlings (Table 2 in paper III). A possible reason for the lower rhizosphere PE of SOM decomposition in the treatment with five pine seedlings is that the root exudation rate expressed as per gram root was reduced with almost 50% compared to pots with one intact pine seedling (Fig. 1 in paper III). However, the results again suggested the variation in the rhizosphere PE on SOM decomposition was more strongly regulated by soil N availability than the root exudation rate.

In paper IV, elevated CO2 increased the root exudation rate per gram root, but only when the elevated CO2 treatment was not combined with N fertilization (Fig. 2 in paper IV). The effect of elevated CO2 on root exudation rates could not explain the detected variation of rhizosphere PEs among treatments. As in paper II and paper III, the results rather suggested that the highly variable rhizosphere PEs was induced by CO2 and N fertilization effects on the partitioning of N between plants and microorganisms, which in turn affected the microbial decomposition of SOM.

In summary, my thesis showed that the root exudation rate is not a decisive factor in regulating the rhizosphere PEs. However, it is also possible that measurements of root exudation rates per gram or cm-3 soil poorly represent the exudate concentration at scales relevant for the microbial community. Soil microbial decomposers respond to conditions in their immediate vicinity (Kuzyakov and Blagodatskaya, 2015), meaning that root exudates expressed on weight or volume basis might not be relevant at the scale experienced by microbial decomposers.

41

Soil C and N availability effect on the depolymerization In Paper V, labile N addition significantly reduced the chitin depolymerization, while did not affect the depolymerization of protein and lignin (Fig. 1 in paper V). Degradation of chitin appears to have provided the microbial community with N under N limited condition, but chitin was replaced by the added N as the microbial N source when N availability increased. Glucose addition stimulated the depolymerization of SOM and the microbial N demand but reduced the depolymerization of lignin, chitin, and protein (Fig. 1 in paper V). I suggested that this was because the labile C input reduced the mineral or physical protection of native SOM compounds, but this protection was not present in the added lignin, chitin, and protein. Hence, labile C input increased the SOM mobilization, which replaced the lignin, chitin, and protein as the microbial nutrient source. Protein and chitin likely account for a large part of the microbial N sources in soil. Therefore, these results contrast the hypothesis that positive PEs are caused by the microbial decomposition of such N rich polymers under N low conditions (Kuzyakov et al., 2000; Dijkstra et al., 2013).

The role of the different microbial groups in the PEs In paper III, I used a combination of 13C and D stable isotope probing methods to investigate the role of different microbial groups in the PE. By quantifying the incorporation of 13C derived from root exudates and D derived from water into microbial biomarker PLFAs, I attempted to separate microbial groups that grew on root exudates or on resources released by SOM decomposition. Microorganisms that increase their growth on SOM derived compounds mobilized by the PE, as indicated by the incorporation of water derived D into their biomass, without relying on 13C from root exudates were assumed to be responsible for the PEs (Hungate et al., 2015). Taken together, the microbial 13C and D incorporation and microbial growth rates detected in the different treatments (Fig. 2 and 3 in paper III), suggested that fungi were responsible for the positive PE in the pine treatment. However, an opportunistic subset of the fungal community that relied on root-derived 13C while depleting available N, resulted in a reduction of microbial SOM decomposer activity in the spruce treatment, resulting in negative PEs.

These results show that different functional groups in the fungal community can induce both positive and negative rhizosphere PE, meaning that measurements of the total fungal community biomass and activity are not sufficient to elucidate their role in SOM decomposition, nutrient cycling, and PEs.

42

The role of the extracellular enzyme in the PEs In paper III, my results showed that the potential activity of extracellular enzymes were all significantly different between the pine and spruce treatments (Fig. 2 in paper III). However, this variation of extracellular enzyme activity alone cannot explain the difference in rhizosphere PEs between the two treatments. My results also suggested that the concentration of C- and N-targeting enzymes were not a limiting factor for SOM decomposition, which might rather have been limited by the activity of oxidative enzymes, such as peroxidases and phenol oxidases. Accordingly, oxidative enzymes that depolymerize macromolecules and make them available for microbial assimilation are of particular importance for the initial stages of SOM decomposition (Hofrichter et al., 2010; Lindahl and Tunlid, 2015). However, since the enzyme activity appears to be more important than the enzyme concentration, which is what the assays I have used measure, the oxidative enzyme concentration could not explain the observed patterns in rhizosphere PEs. IN any case, the weak correlation between the potential enzyme activity and the observed rhizosphere PEs suggest that PEs are not a result of increased extracellular enzyme production.

In paper V, the potential N-targeting enzyme activities were not affected by the C addition and partly affected by the N addition. The potential C-targeting enzyme activities showed different responses to the C addition and generally increased by N addition. The potential oxidative enzyme activity response to C and N addition depended on polymers amendment (Table 2 in paper V). Furthermore, the decomposition of lignin-, chitin-, and protein-derived C was only partly connected to the corresponding potential enzyme activities, and the chitin- and protein-derived C recovered did not correlate with the corresponding hydrolytic enzymes. However, the lignin depolymerization was partly matched by the variation in oxidative enzymes (Fig. 2 in paper V). These results together suggested that oxidative enzyme activities might be more representative of depolymerization reactions than hydrolytic enzyme activities.

In summary, these results again showed that the link between extracellular enzyme activity and SOM decomposition is complicated to elucidate, and suggests that the potential activity of hydrolytic and oxidative enzymes are poor indicators of SOM decomposition and the rhizosphere PE.

43

Conclusion and future perspectives

Rhizosphere interactions are of importance for the release of fixed plant C back to the atmosphere and regulate all aspects of nutrient cycling (Lambers et al., 2008; Stuart et al., 2011; Hopkins et al., 2013). The rhizosphere PE, which is defined as the stimulation or suppression of soil organic matter decomposition in response to labile C input, is increasingly considered as a central role in regulating soil C and N cycling, as well as the atmospheric CO2 concentration (Cheng et al., 2014). However, the underlying mechanisms and controlling factors of rhizosphere PEs are still unclear. In this thesis, I performed a series of experiments to investigate the mechanisms and regulations of the PE on SOM decomposition and gross N mineralization.

In summary, the main findings of my thesis are:

• The magnitude and direction of the rhizosphere PE vary strongly between different tree species.

• Light intensity is an important regulator of the rhizosphere PE, but the effect is plant species dependent.

• The effect of N-fertilization and elevated CO2 on the partitioning of N between plants and microbes is of importance for SOM decomposition and rhizosphere PEs

• SOM decomposition and the PE are strongly regulated by N availability.

• Fungi are more important than bacteria in regulating the PEs.

Taken together, my findings suggest that soil C and N cycling is strongly affected by the PE, which is interactively controlled by multiple regulators. Soil N availability seems to have a particularly strong influence on the PE, due to its concentration-dependent effect on plant and microbial community interactions, which needs to be considered in order to accurately predict PEs. My results also showed that root exudation rate measured on soil weight or volume basis might not be a relevant measure of the conditions at the scale experienced by microorganisms and that this needs to be considered when we investigating how factors that influence the root exudation rate will also affect the PE. My results further highlight the importance of considering the metabolic diversity of particularly the fungal community when assessing or modeling their role in regulating the PE.

44

In conclusion, my studies suggest that PEs can strongly influence SOM decomposition and nutrients cycling. Therefore, the regulation of the PE on SOM decomposition and nutrients cycling could play an important role in the terrestrial ecosystem's C and N budget, and their feedback to global change in response to e.g. elevated CO2 and N deposition.

45

Acknowledgments

First and foremost, I would like to thank my supervisor Per Bengtson. Thank you for giving me the opportunity to do a Ph.D. with you. You have been such an encouraging and perfect supervisor. Thank you for your patience, your support, and invaluable advice over the four years. I could not have asked for a better supervisor.

Johannes Rousk, thanks to being my assistant supervisor. You are always passionate about research. Thanks for your help and guidance in my study in Lund.

Erland Bååth, I cannot study in MEMEG if without your help. You are a learned elder and it is a great pleasure for me that I have the opportunity to learn from you.

Anders Tunlid, thank you for being my examiner and help during my study in MEMEG.

Anna, Edith, Eva, Håkan, Jürgen, and Tomas. You are all very inspiring and nice researchers. Thanks for all your support and help for the last four years.

Saeed Alaei, thank you for all your help in the last four years. I would not accomplish my Ph.D. project without your assistant. I am so happy I can work together with you.

Thanks everyone at Microbial Ecology. Ainara, Annelein, Carlos, Dimitrios, Kristin, Juan, Lettice, Luigi, Margarida, Michiel, Nicholas, Paola, Mingyue. Whenever I asked for help, you would always help me without any hesitation. I am so lucky can work in such a nice working environment.

Baojian Ding, Dandan Zhang, Honglei Wang, Jing Zhang, Hao Zou, Moyan Zhou, Nan Hu, Tianhao Zhao, Tao Wang, Weizhao Yang, Xiaoqing Hou, Xiaozhe Bao, Xiuqing Zhong, Xueying Li, Ye Xiong, Yihan Xia, Yutao Wang, Zhaomo Tian.感谢能在他乡与你们在这里相遇。你们在这几年中给予了我太多太多的帮助，让我体会到了家人的感觉。谢谢你们。我会想念与你们一起的日子。

Chuanshuai Li, Cheng Gong, Dandan Wang, Guangqi Huang, Guangqi Qing, Jiatang Wang, Jian Liu, Jingxue Pan, Haoran Yu, Huayi Gao, Hong Jiang, Lu Yu, Mengshu Hao, Miaoxin Gong, Miao Zhang, Mi Huang, Peibo Li, Qi Jiang, Raojie Lu, Ruoyu Wang, Senbin Yu, Shuo Dong, Suxun Pan, Tian Qi, Wei Wang, Weiming Huang, Xianshao Zou, Xiangkun Yan, Xiaoxu Cui, Xiaoyan Xu, Xingchen Li, Yashu Wang, Yaxiong Deng, Yanzi Yan, Yao Wang, Yichen Liu,

46

Yining Zhu, Yiheng Tong, Yitang Li, Yuxi Zhao, Yunan Zhou, Yong Li, Zhaoyang Liu, Zhengjun Wang and so on~~~能在 Lund 遇到志趣相投的你们真的很庆幸，感谢你们在这几年的生活中对我的各种帮助，我会永远记得我们一起玩耍的日子。

Every life is a story, thank you for being part of my story. Thanks to everyone I have met.

最后，特别感谢我的父母和家人，是你们无私的爱给了我在漫漫求学路上坚持下去的力量与勇气！

47

References

Aciego Pietri, J.C., Brookes, P.C., 2009. Substrate inputs and pH as factors controlling microbial biomass, activity and community structure in an arable soil. Soil Biology and Biochemistry 41, 1396-1405.

Ainsworth, E.A., Long, S.P., 2005. What have we learned from 15 years of free‐air CO2 enrichment (FACE)? A meta‐analytic review of the responses of photosynthesis, canopy properties and plant production to rising CO2. New Phytologist 165, 351-372.

Albiach, R., Canet, R., Pomares, F., Ingelmo, F., 2000. Microbial biomass content and enzymatic activities after the application of organic amendments to a horticultural soil. Bioresource Technology 75, 43-48.

Ander, P., Marzullo, L., 1997. Sugar oxidoreductases and veratryl alcohol oxidase as related to lignin degradation. Journal of Biotechnology 53, 115-131.

Aye, N.S., Butterly, C.R., Sale, P.W.G., Tang, C., 2018. Interactive effects of initial pH and nitrogen status on soil organic carbon priming by glucose and lignocellulose. Soil Biology and Biochemistry 123, 33-44.

Aye, N.S., Butterly, C.R., Sale, P.W.G., Tang, C.X., 2017. Residue addition and liming history interactively enhance mineralization of native organic carbon in acid soils. Biology and Fertility of Soils 53, 61-75.

Bååth, E., 1994. Measurement of protein synthesis by soil bacterial assemblages with the leucine incorporation technique. Biology and Fertility of Soils 17, 147-153.

Bååth, E., 2001. Estimation of fungal growth rates in soil using 14C-acetate incorporation into ergosterol. Soil Biology and Biochemistry 33, 2011-2018.

Bååth, E., Anderson, T.H., 2003. Comparison of soil fungal/bacterial ratios in a pH gradient using physiological and PLFA-based techniques. Soil Biology and Biochemistry 35, 955-963.

Bastida, F., Torres, I.F., Hernandez, T., Bombach, P., Richnow, H.H., Garcia, C., 2013. Can the labile carbon contribute to carbon immobilization in semiarid soils? Priming effects and microbial community dynamics. Soil Biology and Biochemistry 57, 892-902.

Beiler, K., Durall, D., Simard, S., Maxwell, S., Kretzer, A., 2010. Mapping the wood-wide web: mycorrhizal networks link multiple Douglas-fir cohorts. New Phytologist 185, 543-553.

Bell, C., Stromberger, M., Wallenstein, M., 2014. New insights into enzymes in the environment. Biogeochemistry 117, 1-4.

48

Bengtson, P., Barker, J., Grayston, S.J., 2012. Evidence of a strong coupling between root exudation, C and N availability, and stimulated SOM decomposition caused by rhizosphere priming effects. Ecology and Evolution 2, 1843-1852.

Bengtson, P., Bengtsson, G., 2005. Bacterial immobilization and remineralization of N at different growth rates and N concentrations. FEMS Microbiology Ecology 54, 13-19.

Blagodatskaya, E., Kuzyakov, Y., 2008. Mechanisms of real and apparent priming effects and their dependence on soil microbial biomass and community structure: critical review. Biology and Fertility of Soils 45, 115-131.

Blagodatsky, S., Blagodatskaya, E., Yuyukina, T., Kuzyakov, Y., 2010. Model of apparent and real priming effects: Linking microbial activity with soil organic matter decomposition. Soil Biology and Biochemistry 42, 1275-1283.

Burns, R.G., DeForest, J.L., Marxsen, J., Sinsabaugh, R.L., Stromberger, M.E., Wallenstein, M.D., Weintraub, M.N., Zoppini, A., 2013. Soil enzymes in a changing environment: Current knowledge and future directions. Soil Biology and Biochemistry 58, 216-234.

Carreiro, M., Sinsabaugh, R., Repert, D., Parkhurst, D., 2000. Microbial enzyme shifts explain litter decay responses to simulated nitrogen deposition. Ecology 81, 2359-2365.

Chalot, M., Brun, A., 1998. Physiology of organic nitrogen acquisition by ectomycorrhizal fungi and ectomycorrhizas. FEMS Microbiology Reviews 22, 21-44.

Chen, R.R., Senbayram, M., Blagodatsky, S., Myachina, O., Dittert, K., Lin, X.G., Blagodatskaya, E., Kuzyakov, Y., 2014. Soil C and N availability determine the priming effect: microbial N mining and stoichiometric decomposition theories. Global Change Biology 20, 2356-2367.

Cheng, W., Kuzyakov, Y., 2005. Root effects on soil organic matter decomposition. Madison, Wisconsin, USA, Soil Science Society of America.

Cheng, W.X., 1999. Rhizosphere feedbacks in elevated CO2. Tree Physiology 19, 313-320. Cheng, W.X., Johnson, D.W., Fu, S.L., 2003. Rhizosphere effects on decomposition:

Controls of plant species, phenology, and fertilization. Soil Science Society of America Journal 67, 1418-1427.

Cheng, W.X., Parton, W.J., Gonzalez-Meler, M.A., Phillips, R., Asao, S., McNickle, G.G., Brzostek, E., Jastrow, J.D., 2014. Synthesis and modeling perspectives of rhizosphere priming. New Phytologist 201, 31-44.

Cheng, Y., Wang, J., Mary, B., Zhang, J.-b., Cai, Z.-c., Chang, S.X., 2013. Soil pH has contrasting effects on gross and net nitrogen mineralizations in adjacent forest and grassland soils in central Alberta, Canada. Soil Biology and Biochemistry 57, 848-857.

Davidson, E., Hart, S., Shanks, C., Firestone, M., 1991. Measuring gross nitrogen mineralization, and nitrification by 15 N isotopic pool dilution in intact soil cores. Journal of Soil Science 42, 335-349.

Davidson, E.A., Janssens, I.A., 2006. Temperature sensitivity of soil carbon decomposition and feedbacks to climate change. Nature 440, 165-173.

49

Demoling, F., Figueroa, D., Baath, E., 2007. Comparison of factors limiting bacterial growth in different soils. Soil Biology and Biochemistry 39, 2485-2495.

Di Lonardo, D.P., de Boer, W., Zweers, H., van der Wal, A., 2019. Effect of the amount of organic trigger compounds, nitrogen and soil microbial biomass on the magnitude of priming of soil organic matter. PLoS ONE 14.

Dijkstra, F.A., Carrillo, Y., Pendall, E., Morgan, J.A., 2013. Rhizosphere priming: a nutrient perspective. Frontiers in Microbiology 4, 216.

Dijkstra, F.A., Cheng, W., 2007. Moisture modulates rhizosphere effects on C decomposition in two different soil types. Soil Biology and Biochemistry 39, 2264-2274.

Dijkstra, F.A., Cheng, W.X., Johnson, D.W., 2006. Plant biomass influences rhizosphere priming effects on soil organic matter decomposition in two differently managed soils. Soil Biology and Biochemistry 38, 2519-2526.

Dijkstra, F.A., Morgan, J.A., Blumenthal, D., Follett, R.F., 2010. Water limitation and plant inter-specific competition reduce rhizosphere-induced C decomposition and plant N uptake. Soil Biology and Biochemistry 42, 1073-1082.

Fang, Y., Nazaries, L., Singh, B.K., Singh, B.P., 2018. Microbial mechanisms of carbon priming effects revealed during the interaction of crop residue and nutrient inputs in contrasting soils. Global Change Biology 24, 2775-2790.

Fontaine, S., Henault, C., Aamor, A., Bdioui, N., Bloor, J.M.G., Maire, V., Mary, B., Revaillot, S., Maron, P.A., 2011. Fungi mediate long term sequestration of carbon and nitrogen in soil through their priming effect. Soil Biology and Biochemistry 43, 86-96.

Fontaine, S., Mariotti, A., Abbadie, L., 2003. The priming effect of organic matter: a question of microbial competition? Soil Biology and Biochemistry 35, 837-843.

Frostegård, Å., Bååth, E., 1996. The use of phospholipid fatty acid analysis to estimate bacterial and fungal biomass in soil. Biology and Fertility of Soils 22, 59-65.

Fu, S., Cheng, W., 2002. Rhizosphere priming effects on the decomposition of soil organic matter in C4 and C3 grassland soils. Plant and Soil 238, 289-294.

Garcia-Pausas, J., Paterson, E., 2011. Microbial community abundance and structure are determinants of soil organic matter mineralisation in the presence of labile carbon. Soil Biology and Biochemistry 43, 1705-1713.

Godbold, D.L., Hoosbeek, M.R., Lukac, M., Cotrufo, M.F., Janssens, I.A., Ceulemans, R., Polle, A., Velthorst, E.J., Scarascia-Mugnozza, G., De Angelis, P., 2006. Mycorrhizal hyphal turnover as a dominant process for carbon input into soil organic matter. Plant and Soil 281, 15-24.

Gregory, P.J., 2006. Roots, rhizosphere and soil: the route to a better understanding of soil science? European Journal of Soil Science 57, 2-12.

Guenet, B., Camino-Serrano, M., Ciais, P., Tifafi, M., Maignan, F., Soong, J.L., Janssens, I.A., 2018. Impact of priming on global soil carbon stocks. Global Change Biology 24, 1873-1883.

50

Guenet, B., Neill, C., Bardoux, G., Abbadie, L., 2010. Is there a linear relationship between priming effect intensity and the amount of organic matter input? Applied Soil Ecology 46, 436-442.

Hamer, U., Marschner, B., 2005. Priming effects in different soil types induced by fructose, alanine, oxalic acid and catechol additions. Soil Biology and Biochemistry 37, 445-454.

He, P., Wan, S., Fang, X., Wang, F., Chen, F., 2016. Exogenous nutrients and carbon resource change the responses of soil organic matter decomposition and nitrogen immobilization to nitrogen deposition. Scientific Reports 6, 23717.

Heimann, M., Reichstein, M., 2008. Terrestrial ecosystem carbon dynamics and climate feedbacks. Nature 451, 289-292.

Hicks, L.C., Meir, P., Nottingham, A.T., Reay, D.S., Stott, A.W., Salinas, N., Whitaker, J., 2018. Carbon and nitrogen inputs differentially affect priming of soil organic matter in tropical lowland and montane soils. Soil Biology and Biochemistry 129, 212-222.

Hobbie, J.E., Hobbie, E.A., 2013. Microbes in nature are limited by carbon and energy: the starving-survival lifestyle in soil and consequences for estimating microbial rates. Frontiers in Microbiology 4, 324.

Hofrichter, M., Ullrich, R., Pecyna, M.J., Liers, C., Lundell, T., 2010. New and classic families of secreted fungal heme peroxidases. Applied Microbiology and Biotechnology 87, 871-897.

Hopkins, F., Gonzalez‐Meler, M.A., Flower, C.E., Lynch, D.J., Czimczik, C., Tang, J., Subke, J.A., 2013. Ecosystem‐level controls on root‐rhizosphere respiration. New Phytologist 199, 339-351.

Hungate, B.A., Mau, R.L., Schwartz, E., Caporaso, J.G., Dijkstra, P., van Gestel, N., Koch, B.J., Liu, C.M., McHugh, T.A., Marks, J.C., Morrissey, E.M., Price, L.B., 2015. Quantitative Microbial Ecology through Stable Isotope Probing. Applied and Environmental Microbiology 81, 7570-7581.

Huo, C.F., Luo, Y.Q., Cheng, W.X., 2017. Rhizosphere priming effect: A meta-analysis. Soil Biology and Biochemistry 111, 78-84.

IAEA, 2001. Use of isotope and radiation methods in soil and water management and crop nutrition. International Atomic Energy Agency, Vienna, Austria.

Jones, D., Darrah, P., 1993. Re-sorption of organic compounds by roots of Zea mays L. and its consequences in the rhizosphere. Plant and Soil 153, 47-59.

Kemmitt, S.J., Wright, D., Goulding, K.W.T., Jones, D.L., 2006. pH regulation of carbon and nitrogen dynamics in two agricultural soils. Soil Biology and Biochemistry 38, 898-911.

Knicker, H., 2011. Soil organic N-An under-rated player for C sequestration in soils? Soil Biology and Biochemistry 43, 1118-1129.

Kumar, A., Kuzyakov, Y., Pausch, J., 2016. Maize rhizosphere priming: field estimates using 13C natural abundance. Plant and Soil 409, 87-97.

Kurets, V., Popov, E., Talanov, A., Drozdov, S., 1994. Light and Temperature Characteristics of CO2 Exchange in Pine and Spruce Seedlings. Russian Journal of Plant Physiology 41, 468-473.

51

Kuzyakov, Y., 2002. Factors affecting rhizosphere priming effects. Journal of Plant Nutrition and Soil Science 165, 382-396.

Kuzyakov, Y., Blagodatskaya, E., 2015. Microbial hotspots and hot moments in soil: concept and review. Soil Biology and Biochemistry 83, 184-199.

Kuzyakov, Y., Cheng, W., 2001. Photosynthesis controls of rhizosphere respiration and organic matter decomposition. Soil Biology and Biochemistry 33, 1915-1925.

Kuzyakov, Y., Ehrensberger, H., Stahr, K., 2001. Carbon partitioning and below-ground translocation by Lolium perenne. Soil Biology and Biochemistry 33, 61-74.

Kuzyakov, Y., Friedel, J.K., Stahr, K., 2000. Review of mechanisms and quantification of priming effects. Soil Biology and Biochemistry 32, 1485-1498.

Kuzyakov, Y., Xu, X.L., 2013. Competition between roots and microorganisms for nitrogen: mechanisms and ecological relevance. New Phytologist 198, 656-669.

Lal, R., 2009. Challenges and opportunities in soil organic matter research. European Journal of Soil Science 60, 158-169.

Lal, R., 2018. Digging deeper: A holistic perspective of factors affecting soil organic carbon sequestration in agroecosystems. Global Change Biology 24, 3285-3301.

Lambers, H., Chapin III, F.S., Pons, T.L., 2008. Plant physiological ecology. Springer Science and Business Media.

Li, L., Zhu-Barker, X., Ye, R., Doane, T.A., Horwath, W.R., 2018. Soil microbial biomass size and soil carbon influence the priming effect from carbon inputs depending on nitrogen availability. Soil Biology and Biochemistry 119, 41-49.

Lindahl, B.D., Tunlid, A., 2015. Ectomycorrhizal fungi - potential organic matter decomposers, yet not saprotrophs. New Phytologist 205, 1443-1447.

Lu, J.Y., Dijkstra, F.A., Wang, P., Cheng, W.X., 2018. Rhizosphere priming of grassland species under different water and nitrogen conditions: a mechanistic hypothesis of C-N interactions. Plant and Soil 429, 303-319.

Månsson, K., Bengtson, P., Falkengren‐Grerup, U., Bengtsson, G., 2009. Plant–microbial competition for nitrogen uncoupled from soil C: N ratios. Oikos 118, 1908-1916.

Marx, M.C., Wood, M., Jarvis, S.C., 2001. A microplate fluorimetric assay for the study of enzyme diversity in soils. Soil Biology and Biochemistry 33, 1633-1640.

Mary, B., Recous, S., Robin, D., 1998. A model for calculating nitrogen fluxes in soil using 15N tracing. Soil Biology and Biochemistry 30, 1963-1979.

Merilä, P., Derome, J., 2008. Relationships between needle nutrient composition in Scots pine and Norway spruce stands and the respective concentrations in the organic layer and in percolation water. Boreal Environment Research 13, 35-47.

Otten, W., Hall, D., Harris, K., Ritz, K., Young, I.M., Gilligan, C.A., 2001. Soil physics, fungal epidemiology and the spread of Rhizoctonia solani. New Phytologist 151, 459-468.

Payton, M.E., Greenstone, M.H., Schenker, N., 2003. Overlapping confidence intervals or standard error intervals: what do they mean in terms of statistical significance? Journal of Insect Science 3, 34.

52

Payton, M.E., Miller, A.E., Raun, W.R., 2000. Testing statistical hypotheses using standard error bars and confidence intervals. Communications in Soil Science and Plant Analysis 31, 547-551.

Phillips, R.P., Fahey, T.J., 2006. Tree species and mycorrhizal associations influence the magnitude of rhizosphere effects. Ecology 87, 1302-1313.

Qiao, N., Xu, X., Hu, Y., Blagodatskaya, E., Liu, Y., Schaefer, D., Kuzyakov, Y., 2016. Carbon and nitrogen additions induce distinct priming effects along an organic-matter decay continuum. Scientific Reports 6, 19865.

Rousk, J., Brookes, P.C., Bååth, E., 2009. Contrasting soil pH effects on fungal and bacterial growth suggest functional redundancy in carbon mineralization. Applied and Environmental Microbiology 75, 1589-1596.

Rousk, J., Hill, P.W., Jones, D.L., 2015. Priming of the decomposition of ageing soil organic matter: concentration dependence and microbial control. Functional Ecology 29, 285-296.

Saiya-Cork, K.R., Sinsabaugh, R.L., Zak, D.R., 2002. The effects of long term nitrogen deposition on extracellular enzyme activity in an Acer saccharum forest soil. Soil Biology and Biochemistry 34, 1309-1315.

Schimel, J.P., Bennett, J., 2004. Nitrogen mineralization: challenges of a changing paradigm. Ecology 85, 591-602.

Schmidt, M.W.I., Torn, M.S., Abiven, S., Dittmar, T., Guggenberger, G., Janssens, I.A., Kleber, M., Kogel-Knabner, I., Lehmann, J., Manning, D.A.C., Nannipieri, P., Rasse, D.P., Weiner, S., Trumbore, S.E., 2011. Persistence of soil organic matter as an ecosystem property. Nature 478, 49-56.

Simard, S.W., Perry, D.A., Jones, M.D., Myrold, D.D., Durall, D.M., Molina, R., 1997. Net transfer of carbon between ectomycorrhizal tree species in the field. Nature 388, 579-582.

Sinsabaugh, R.L., Lauber, C.L., Weintraub, M.N., Ahmed, B., Allison, S.D., Crenshaw, C., Contosta, A.R., Cusack, D., Frey, S., Gallo, M.E., Gartner, T.B., Hobbie, S.E., Holland, K., Keeler, B.L., Powers, J.S., Stursova, M., Takacs-Vesbach, C., Waldrop, M.P., Wallenstein, M.D., Zak, D.R., Zeglin, L.H., 2008. Stoichiometry of soil enzyme activity at global scale. Ecology Letters 11, 1252-1264.

Sterner, R.W., Elser, J.J., 2002. Ecological stoichiometry: the biology of elements from molecules to the biosphere. Princeton university press.

Strickland, M.S., Lauber, C., Fierer, N., Bradford, M.A., 2009. Testing the functional significance of microbial community composition. Ecology 90, 441-451.

Stuart, F., Matson, P.A., Vitousek, P., 2011. Principles of terrestrial ecosystem ecology. Springer Science and Business Media.

Su, P., Lou, J., Brookes, P.C., Luo, Y., He, Y., Xu, J.M., 2017. Taxon-specific responses of soil microbial communities to different soil priming effects induced by addition of plant residues and their biochars. Journal of Soils and Sediments 17, 674-684.

Teste, F.P., Simard, S.W., Durall, D.M., Guy, R.D., Berch, S.M., 2010. Net carbon transfer between Pseudotsuga menziesii var. glauca seedlings in the field is influenced by soil disturbance. Journal of Ecology 98, 429-439.

53

Thiessen, S., Gleixner, G., Wutzler, T., Reichstein, M., 2013. Both priming and temperature sensitivity of soil organic matter decomposition depend on microbial biomass–An incubation study. Soil Biology and Biochemistry 57, 739-748.

Tiedje, J.M., 1982. Denitrification, Methods of soil analysis. Agronomy Monogr, Madison, Wisc (1982), pp. 1011-1026.

Van Hees, P.A., Jones, D.L., Finlay, R., Godbold, D.L., Lundström, U.S., 2005. The carbon we do not see—the impact of low molecular weight compounds on carbon dynamics and respiration in forest soils: a review. Soil Biology and Biochemistry 37, 1-13.

Wang, H., Xu, W.H., Hu, G.Q., Dai, W.W., Jiang, P., Bai, E., 2015. The priming effect of soluble carbon inputs in organic and mineral soils from a temperate forest. Oecologia 178, 1239-1250.

Wang, J.Y., Dokohely, M.E., Xiong, Z.Q., Kuzyakov, Y., 2016a. Contrasting effects of aged and fresh biochars on glucose-induced priming and microbial activities in paddy soil. Journal of Soils and Sediments 16, 191-203.

Wang, X.J., Tang, C.X., Severi, J., Butterly, C.R., Baldock, J.A., 2016b. Rhizosphere priming effect on soil organic carbon decomposition under plant species differing in soil acidification and root exudation. New Phytologist 211, 864-873.

Warembourg, F.R., Estelrich, H.D., 2001. Plant phenology and soil fertility effects on below-ground carbon allocation for an annual (Bromus madritensis) and a perennial (Bromus erectus) grass species. Soil Biology and Biochemistry 33, 1291-1303.

Wild, B., Alaei, S., Bengtson, P., Bodé, S., Boeckx, P., Schnecker, J., Mayerhofer, W., Rütting, T., 2017. Short-term carbon input increases microbial nitrogen demand, but not microbial nitrogen mining, in a set of boreal forest soils. Biogeochemistry 136, 261-278.

Xiao, C.W., Guenet, B., Zhou, Y., Su, J.Q., Janssens, I.A., 2015. Priming of soil organic matter decomposition scales linearly with microbial biomass response to litter input in steppe vegetation. Oikos 124, 649-657.

Xu, Q., Wang, X., Tang, C., 2017. Wheat and white lupin differ in rhizosphere priming of soil organic carbon under elevated CO2. Plant and Soil 421, 43-55.

Xu, Q., Wang, X., Tang, C., 2018. The effects of elevated CO2 and nitrogen availability on rhizosphere priming of soil organic matter under wheat and white lupin. Plant and Soil 425, 375-387.

Yin, L., Dijkstra, F.A., Wang, P., Zhu, B., Cheng, W., 2018. Rhizosphere priming effects on soil carbon and nitrogen dynamics among tree species with and without intraspecific competition. New Phytologist 218, 1036-1048.

Yiqi, L., Zhou, X., 2010. Soil respiration and the environment. Elsevier, San Diego, CA, USA. .

Zak, D.R., Pellitier, P.T., Argiroff, W., Castillo, B., James, T.Y., Nave, L.E., Averill, C., Beidler, K.V., Bhatnagar, J., Blesh, J., 2019. Exploring the role of ectomycorrhizal fungi in soil carbon dynamics. New Phytologist 223, 33-39.

Zhao, Z., Ge, T., Gunina, A., Li, Y., Zhu, Z., Peng, P., Wu, J., Kuzyakov, Y., 2018. Carbon and nitrogen availability in paddy soil affects rice photosynthate allocation,

54

microbial community composition, and priming: combining continuous 13C labeling with PLFA analysis. Plant and Soil.

Zhu, B., Gutknecht, J.L.M., Herman, D.J., Keck, D.C., Firestone, M.K., Cheng, W.X., 2014. Rhizosphere priming effects on soil carbon and nitrogen mineralization. Soil Biology and Biochemistry 76, 183-192.

JIAN

LI

Mechanism

s and regulations of priming effects in soil

2019

Department of BiologyFaculty of Science

Lund University, ISBN 978-91-7895-360-8

Mechanisms and regulations of priming effects in soilJIAN LI

FACULTY OF SCIENCE | LUND UNIVERSITY

List of papers






9789178

953608

mechanisms and regulations of priming effects in soil li, jian

Documents