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Biochar effects on soil nutrient transformations
Chapter · January 2015
4 authors, including:
Thomas H Deluca
M Derek Mackenzie
University of Montana
University of Alberta
Davey L Jones
University of Western Australia
Some of the authors of this publication are also working on these related projects:
'CINAg': UK-China Virtual Joint Centre for Improved Nitrogen Agronomy View project
Soil and Foliar Chemistry Analysis for WBEA’s Terrestrial Environmental Effects Monitoring View project
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Biochar effects on soil nutrient transformations
Thomas H. DeLuca, Michael J. Gundale, M. Derek MacKenzie
and Davey L. Jones
A key aspect of biochar management is its
proposed effect on soil fertility and plant production (Lehmann, 2007). Plant productivity
is directly influenced by nutrient availability,
which is a product of nutrient transformations in the soil environment (Marschner,
2006). For that reason, there is a great deal of
interest in how biochar influences nutrient
transformations and availability of those
nutrients for plant uptake. And although
increasing evidence suggests that biochar
addition to soil may enhance plant production in a variety of natural and agricultural
environments (Lehmann and Rondon, 2006;
Atkinson et al, 2010; Jeffery et al, 2011), the
direct influence of biochar additions on soil
nutrient cycling are inconsistent and remain
somewhat poorly understood. This is partially due to differences in the soils, crops and
biochar amendments used in experiments
and the predominance of short-term studies
in the literature.
The purpose of this chapter is to summarize several general mechanisms through
Biochar C15.indd 419
which biochar affects nutrient availability to
plants and to specifically evaluate the effect
biochar has on nutrient cycling and specific
transformations within several key nutrient
cycles. Our aim is to complement other chapters in this book as well as other reviews that
have focused on the chemical and physical
properties of biochar (Chapters 5 and 6)
(Atkinson et al, 2010; Shrestha et al, 2010),
its stability in the soil (Shrestha et al, 2010),
its effect on soil biota and plant eco-physiology (Chapters 13 and 14) (Atkinson et al,
2010; Lehmann et al, 2011), its influence on
soil properties (Atkinson et al, 2010; Beesley
et al, 2011), its impact on trace gas fluxes
(Chapter 18) (Jones et al, 2011) and its use in
biofuel technology or soil C sequestration
(Chapter 10) (Lehmann, 2007; Lee et al,
2010; Macias and Arbestain, 2010). The
purpose of this review is to explore what is
and is not known about the effect of biochar
on soil nutrient transformations, which are
likely to have both short- and long-term
impacts on plant productivity in forest and
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agricultural landscapes (Atkinson et al, 2010).
We specifically focus on the response of
nitrogen (N), phosphorus (P) and sulfur (S)
cycles in response to biochar amendment and
explore the implications for altered transformations within these cycles on the availability
of nutrients to plants and their long term
budgets across a range of ecosystems. In this
effort we will try to differentiate between
short-term and long-term effects of biochar
on ecosystem processes.
General mechanisms by which biochar influences nutrient
turnover and transformations
The application of biochar to agricultural and
forest soils has been found to increase the
bio-availability and uptake of many nutrients
to plants (Glaser et al, 2002; Lehmann et al,
2003; Steiner et al, 2007; Jeffery et al, 2011;
Nelson et al, 2011). While some mechanisms
causing increased nutrient availability have
been extensively described and summarized
(Atkinson et al, 2010), very little research has
been conducted on the influence of biochar
on specific nutrient cycling mechanisms. For
instance, numerous studies have described
high concentrations of available nutrients on
the surface of newly created biochar made
over a wide range of temperatures and oxidation conditions, and from a range of feedstocks, suggesting that biochars themselves
can have fertilization effects over short timescales (Chan and Xu, 2009; Jeffery et al,
2011). As an example, the direct contribution
of NH4+ salts from newly formed biochar has
been described in numerous studies (see
Chapter 7; Gundale and DeLuca, 2006;
Chan and Xu, 2009; Spokas et al, 2012).
Considerably less attention has been given to
the effect of biochar on specific transformations, i.e. indirect alterations, within the N
cycle (Lehmann, 2007; Clough and Condron,
2010; Clough et al, 2013), for example biological N-fixation, N-mineralization, nitrification and gaseous N losses (Clough and
Condron, 2010).
The effects that biochar may have on
nutrient transformations has consequences
Biochar C15.indd 420
for the long term effect of biochar on plant
productivity and on nutrient stocks (i.e. the
installation of a positive reinforcing feedback
loop, see Figure 15.1) and therefore have
important implications for the viability and
sustainability of biochar as a C mitigation
strategy (Lehmann, 2007; Roberts et al,
2010). In the following section we identify
three general mechanisms through which biochar may influence nutrient cycles: (i)
Increase in the pool size and turnover of labile
organic nutrients; (ii) Alteration of soil physical and chemical properties; and (iii) effects
on soil biota.
Increase in the pool size and
turnover of labile organic
A primary mechanism by which biochar may
accelerate nutrient cycling over long time
scales is by serving as a short-term source of
highly available nutrients, which become
incorporated into living biomass and labile
soil organic pools (Jeffery et al, 2011). As
described above and in detail in Chapter 7,
new, unweathered biochar, especially that
generated from nutrient-rich material, can be
a source of highly available nutrient salts that
provide a direct short-term source of nutrition to plants (Chan and Xu, 2009; Atkinson
et al, 2010). During the pyrolysis process,
heating causes some nutrients to volatilize
(e.g. N as nitrogen oxides (NOx)), especially
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Figure 15.1 A conceptual model for the
influence of biochar on nutrient (Nr) turnover in
the soil environment
at the surface of the material, while other
nutrients become concentrated in the remaining biochar (Gundale and DeLuca, 2006;
Nelson et al, 2011). Feedstock, formation
temperature, the time a material is held at a
given temperature, oxygen availability and
the heating rate directly influence the surface
residue chemistry of biochar (Gundale and
DeLuca, 2006; Atkinson et al, 2010). Some
specific elements are disproportionately lost
to the atmosphere, fixed into recalcitrant
forms or liberated as soluble oxides during
the heating process, affecting the chemical
composition of ash residues on the biochar
surface (Chan and Xu, 2009). For woodderived biochars, carbon (C) begins to volatilize around 100ºC, N above 200°C, S above
Biochar C15.indd 421
375 ºC and potassium (K) and P between
700 and 800ºC (Neary et al, 2005), whereas
the volatilization of magnesium (Mg), calcium (Ca) and manganese (Mn) only occurs
at temperatures above 1000ºC (Neary et al,
1999; Knoepp et al, 2005). These differences
in volatilization temperatures among elements cause shifts in the stoichiometry of biochar elemental concentrations, with total S
and N concentrations often decreasing relative to other elements due to their lower volatilization temperatures (Knudsen et al, 2004;
Trompowsky et al, 2005). Correspondingly,
several nutrient salts accumulate on biochar
surfaces, with NH4+ and SO4-2 concentrations
increasing in low temperature biochars (<
500ºC) (Knudsen et al, 2004; Gundale and
DeLuca, 2006) and NO3-, PO4-3, Ca+2, Mg+2
and trace metals increasing, especially in biochars formed at high temperatures (Gundale
and DeLuca, 2006; Chan and Xu, 2009;
Atkinson et al, 2010; Nelson et al, 2011).
Because soils generally contain a relatively large total pool of most nutrients, biochar additions to soil usually provide only a
modest contribution to the total soil nutrient
capital (Chan et al, 2007). However, only a
small fraction of the total soil nutrient capital
is usually bio-available, meaning that the
addition of nutrient salts in biochar surface
residues can constitute a significant increase
in the bio-available pool of some nutrients
(Gundale and DeLuca, 2006; Yamato et al,
2006; Chan et al, 2007). This short-term
input of bio-available nutrients can enhance
plant productivity (i.e. total biomass) and
improve tissue quality (Kimetu et al, 2008;
Jeffery et al, 2011) and therefore influence
both the quantity and quality of nutrient-containing plant residues returned to the soil
(Major et al, 2010). Plants’ C inputs to the
soil occur through root exudation and turnover, and through senescence and death of
above-ground tissues. It is well known that
the nutrient concentrations of plant litter is a
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strong control on nutrient mineralization
rates (Stevenson and Cole, 1999; Brady and
Weil, 2002). Therefore, larger inputs of
higher quality plant organic matter to the soil
in response to biochar derived nutrients likely
result in an increase in the labile nutrient pool
size, thereby in theory increasing the total
quantity of labile organic nutrients returned
back to the soil and available for mineralization (DeLuca et al, 2006; Gundale and
DeLuca, 2007; Major et al, 2010). This feedback involving higher plant nutrient uptake, a
higher return of labile organic nutrients to the
soil and higher nutrient mineralization rates
could enhance nutrient availability to plants
over longer time scales as implied in Figure
15.1. The persistence of accelerated nutrient
turnover between plants and soil is likely
dependent on the size of the nutrient pool
added from biochar, the frequency of its
addition (e.g. single dose, multiple doses or
annual), the degree to which nutrient capital
is removed from a system during harvesting
activities (Jeffery et al, 2011), the degree to
which nutrients are fixed into recalcitrant
organic or insoluble mineral pools (Hedley et
al, 1982), the long-term losses in nutrient
capital through leaching or volatilization that
occur at a given site (Lehmann et al, 2003;
Yanai et al, 2007) and the long-term build-up
(or decline) of stable, recalcitrant organomineral complexes beyond the pure biochar.
Alteration of soil physical and
chemical properties
In addition to its direct contribution of available nutrients to the soil, biochar has a variety
of physical and chemical properties that influence soil nutrient transformations. For a more
detailed review of biochar physical and chemical properties, see Chapters 5 and 6, and
Atkinson et al (2010). Biochar is a high surface area (Beesley et al, 2011), highly porous
(Keech et al, 2005), variable charge organic
Biochar C15.indd 422
material which often contains a surface residue enriched in alkaline metals (Figure 15.2b)
(Atkinson et al, 2010). When added to soil,
biochar has the potential to alter the physical
and chemical properties of soil, which in turn
can influence nutrient transformation rates.
For instance, biochar has been shown to
increase soil water holding capacity, alter gas
exchange, increase cation exchange capacity
(CEC), increase surface sorption capacity,
increase base saturation of acidic mineral
soils and alter soil pH (Glaser et al, 2002;
Bélanger et al, 2004; Keech et al, 2005; Liang
et al, 2006; Atkinson et al, 2010). These biochar properties are highly dependent on
the temperature and duration of pyrolysis
(Gundale and DeLuca, 2006; Bornermann et
al, 2007; Joseph et al, 2007) and the feedstock
from which biochar is made (Gundale and
DeLuca, 2006; Streubel et al, 2011).
Many nutrient transformations are
enzyme mediated reactions carried out by
soil micro-organisms. Soil micro-organisms
require environments with appropriate water
potential and redox conditions to carry out
their metabolic activities (Alexander, 1991;
Briones, 2012). The physical structure of
biochar contains a range of pore sizes which
are inherited from the feedstock material
(Keech et al, 2005) and which can directly
influence the water potential and redox environment of soil micro-organisms (Joseph et
al, 2010). Micropores, defined by soil scientists as pores with < 30µm diameter (Brady
and Weil, 2002), serve as capillary spaces
with high surface area to volume ratios and
can retain water even when soil moisture is
strongly depleted (Kammann et al, 2011),
thereby creating moist microsites (Lehmann
and Rondon, 2006). Biochar also often contains macropores (>75µm diameter) which
can serve as gas exchange channels (Keech et
al, 2005), thereby influencing the redox environment for soil biota (Joseph et al, 2010;
Lehmann et al, 2011). Organic residues
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decompose much more rapidly under aerobic
conditions and therefore biochar may
enhance nutrient mineralization in soils with
inherently poor gas exchange properties
(Gundale and DeLuca, 2006; Asai et al,
2009). Likewise, several specific nutrient
transformations require oxygen as an electron
acceptor, such as nitrification and S oxidation, which suggests that the physical structure of biochar may increase oxidative
transformations in soils with inherently poor
gas exchange environments (DeLuca et al,
2006; Asai et al, 2009; Joseph et al, 2010).
The highly variable pore size distribution of
biochar thus assures the presence of a wide
variety of soil micro-sites with contrasting
moisture and redox conditions under variable
environmental conditions (Joseph et al,
2010). In all biological living systems from
microorganisms to plants, gradients and
potential-differences (concentrations, redox,
Eh, pH) are the prerequisites for material and
electron flow, i.e. for metabolism and energy
gain. The addition of biochars may thus
intensify microbial or root-associated gross
nutrient cycling processes by: (i) creating
more ‘micro-site opportunities’ with steeper
redox, pH or nutrient concentration gradients around or across biochar particles
(Briones, 2012; Joseph et al, 2013); and by
(ii) creating more such micro-scale metabolism opportunities within the soil matrix. If
the ‘biochar micro-site metabolism opportunities’ then meet higher organic (e.g. crop or
root residue) inputs, a positive feedback cycle
with intensified gross nutrient cycling may
result, improving soil fertility in the long run
(Figure 15.1).
Additional mechanisms through which
biochar amendments can alter nutrient transformations is by: (i) Reducing nutrient loss
from the soil (Crutchfield et al, 2010; Ding et
al, 2010; Prendergast-Miller et al, 2011;
Ventura et al, 2013); (ii) Reducing fixation of
nutrients into insoluble mineral or recalci-
Biochar C15.indd 423
trant organic pools (Cui et al, 2011; Nelson et
al, 2011); (iii) Reduced gaseous losses of N,
be it as NH3, N2 or N2O (Prendergast-Miller
et al, 2011; Taghizadeh-Toosi et al, 2012a;
Spokas et al, 2012); (iv) By ameliorating
other constraints of nutrient cycling e.g. in
contaminated soils by its sorptive properties
(Figure 15.1). Biochar has been shown to
have a transient anion exchange capacity and
moderately high CEC that also changes with
time in the soil (Brewer et al, 2011). Biochar
also can ameliorate soil pH due to its surface
ash residue containing alkaline metals, especially in acidic soils (Brewer et al, 2011).
Some biochars can harbour relatively high
exchange capacity per unit mass (Atkinson et
al, 2010), therefore its addition to some soils
can increase surface soil exchange capacity.
Free nutrient ions can be lost from ecosystems via leaching or runoff, resulting in a
decline in soil fertility through time. Biochar
can reduce leaching and volatilization losses
under some circumstances (PrendergastMiller et al, 2011; Taghizadeh-Toosi et al,
2012a; Spokas et al, 2012; Ventura et al,
2013) and thus increase the total pool of
nutrients available to plants or microbes. A
variety of studies suggests that biochar can
simultaneously reduce nutrient leaching and
volatilization losses through its influence on
soil pH (Bélanger et al., 2004, Atkinson et al.,
2010) and CEC (Crutchfield et al, 2010;
Ding et al, 2010); however, the alkaline
nature of some biochars may actually increase
NH3 volatilization in surface soils amended
with biochar (Chen et al, 2013).
An additional characteristic of biochar
that can influence nutrient cycling is its effect
on soil solution C chemistry (Figure 15.1) and
turnover (see Chapter 16). While biochar itself
has been shown to contain only a minor fraction of bio-available C (Jones et al, 2010;
Major et al, 2010), several studies suggest that
it serves as a strong sorptive surface for a wide
range of C compounds. The high surface area,
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porous (Figures 15.2 and 15.3) and often
hydrophobic nature of biochar directly after
production makes it an ideal surface for the
sorption of hydrophobic organic compounds
(Cornelissen et al, 2004; Keech et al, 2005;
Bornermann et al, 2007; Gundale and
DeLuca, 2007). Numerous studies have
shown a reduction in soluble or free phenolic
compounds when activated C is added to soils
(DeLuca et al, 2002; Wallstedt et al, 2002;
Berglund et al, 2004; Keech et al, 2005;
Gundale and DeLuca, 2006; MacKenzie and
DeLuca, 2006). Additional studies have demonstrated that char formed during wildfires or
burning of agricultural residues also functions
to adsorb phenolic and various aromatic and
hydrophobic organic compounds (Yaning and
Sheng, 2003; Brimmer, 2006; DeLuca et al,
2006; Gundale and DeLuca, 2006; MacKenzie
and DeLuca, 2006; Bornermann et al, 2007).
Through these sorption reactions, biochar
may: (i) reduce the activity of compounds that
may be either inhibitory to nutrient transformation specialists, such as nitrifying bacteria
(White, 1991; Ward et al, 1997; Paavolainen
et al, 1998); (ii) reduce complexation of nutrient-rich molecules such as proteins into tannin-complexes (Kraus et al, 2003; Gundale et
al, 2010); or (iii) reduce the concentration of
bio-available C in the soil solution that would
otherwise enhance the immobilization of inorganic N, P or S (Schimel et al, 1996; Stevenson
and Cole, 1999) (Figure 15.1). The interaction of soluble soil C with biochar surfaces is a
key mechanism that may influence nutrient
availability and transformations, but has
received relatively little attention (MacKenzie
and DeLuca, 2006; Nelissen et al, 2012).
Alteration of microbial
Biochar additions to soil have the potential to
change the biomass, community composition
and activity of soil microbes, all of which can
Biochar C15.indd 424
Figure 15.2 Electron micrographs of a high
sorption (a) and low sorption (b) char collected
from forest soils in northern Idaho (Brimmer,
2006). The high sorption char (immature char
formed in a recent fire) has open pores that
follow trachieds whereas the low sorption char
(mature char) has many of the pores occluded
with organics
influence nutrient mineralization from
decomposing plant residues, as well as several
specific nutrient transformations. For a complete review of biochar effects on soil microbial communities, see Chapter 13. There are
several mechanisms proposed by which biochar can influence soil microbes, including
the porous structure of biochar which provides a habitat for microbes (Pietikäinen and
Fritze, 1993), its effects on plant growth and
associated plant C inputs (Major et al, 2010),
its source of trace minerals (Rondon et al,
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Figure 15.3 The pH, electrical conductivity (EC), cation exchange capacity (CEC) and density of
biochar produced from Douglas-fir or ponderosa pine wood or bark at 350 or 800°C (from Gundale
and DeLuca, 2006). Data meeting the assumptions of normality were compared with one-way
ANOVA followed by the Student-Neuman-Kuels post hoc procedure where letters indicate pairwise
differences. Non-normal data were compared using the Kruskal-Wallis (K-W) statistic
2007), its sorption of microbial signalling
compounds or inhibitory plant phenolic compounds (DeLuca et al, 2006; Ni et al, 2010),
as well as its effect on soil physical and chemical properties (described on page 422). Very
few studies have attempted to isolate the relative importance of these factors and substantial uncertainty remains regarding the
mechanisms through which biochar influences soil microbial community properties
(Lehmann et al, 2011).
Despite mechanistic uncertainty, several
studies have shown that increases in micro-
Biochar C15.indd 425
bial biomass appear to occur in response to
soil biochar amendments. O’Neill et al (2009)
and Liang et al (2010) showed that
Amazonian Anthrosol soils rich in pyrogenic
C (PyC) contained between one and two
orders of magnitude greater microbial biomass than adjacent soils with low PyC concentrations. In short-term studies with
biochar amendments, Anderson et al (2011)
showed somewhat more modest increases of
between 5–14 per cent in the abundance of a
variety of microbial taxonomic groups in
response to biochar amendment in a temper-
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ate pasture soil. Pietikäinen et al (2000) also
showed that microbial respiration and biomass increased in response to additions of
pyrogenic carbonaceous matter (PCM) in
boreal soils. However, other studies have
shown no significant shift in microbial activity with biochar amendments to soils (Jones et
al, 2012). Although soil microbes are the primary driver of organic nutrient mineralization and oxidative or reductive nutrient
transformations, these studies suggest that
biochar-induced changes in microbial communities likely have consequences for nutrient turnover rates between plants and soil.
In addition to observed shifts in microbial biomass in response to biochar, a variety
of studies have shown that microbial community composition can be altered by biochar
(Anderson et al, 2012; Jones et al, 2012;
Ducey et al, 2013), sometimes resulting in
increased abundance of functional groups
that have key roles in nutrient cycling and
plant nutrient acquisition (Lehmann et al,
2011). Mycorrhizal fungi, which play a key
role in extracting nutrients from recalcitrant
organic or insoluble mineral pools, have been
observed to both increase (Saito, 1990;
Makoto et al, 2010; Solaiman et al, 2010) and
decrease (Warnock et al, 2007; Lehmann et
al, 2011) in response to biochar addition to
soil. Given the specific functional role of
mycorrhizae in nutrient acquisition, changes
in mycorrhizal biomass and colonization
likely influence the flux of nutrients from
unavailable nutrient pools (i.e. recalcitrant
organic matter and insoluble minerals, in particular P) into biomass and therefore labile
organic pools that actively turn over between
plants and soil, although studies identifying
the effect of biochar on these fluxes are lacking. In addition to mycorrhizae, several specific nutrient transformations have been
shown to either increase or decrease in
response to soil biochar amendments and in
some cases altered transformation rates have
been linked to changes in the abundance of
specific soil biota. One of the clearest examples of this is the frequently observed increase
of nitrification rates in biochar amended forest soils (DeLuca et al, 2006; Gundale and
DeLuca, 2007), which has been linked to
larger populations of nitrifying bacteria in
biochar pore spaces (Ball et al, 2010;
described in further detail below). While an
increasing number of studies have described
changes in either microbial community composition or changes in nutrient transformation rates in response to biochar amendment
(Lehmann et al, 2011), there remain a relatively small number that explicitly link altered
nutrient transformation rates to functional
changes in the soil microbial community (Ball
et al, 2010; Ducey et al, 2013).
Influences of biochar on specific nutrient transformations
As previously described, there are a range of
mechanisms through which biochar can
influence the loss of nutrients from forest or
agricultural ecosystems, as well as the gross
annual turnover between soils and plants.
However, plant nutrients differ substantially
in many key physical and chemical properties, such as the number of molecular species
that exist, the mass and potential energy of
Biochar C15.indd 426
those molecular species and the ability of
microbes to transform molecular species to
acquire energy under given soil conditions
(i.e. pH and redox potential). These individualistic properties of different nutrient elements can determine how the cycles of
specific elements respond to soil biochar
amendments. In the following sections we
review the influence of biochar on N, P and S
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cycles. Biochar always contains some quantity of soluble inorganic N and P (see Chapter
7 and Figure 15.4) which can be released rapidly or slowly into the soil; however, in this
section we will focus on the influence of biochar on transformations of nutrients in the
soil as opposed to nutrient delivery.
Nitrogen is the single most limiting plant nutrient in most cold or temperate terrestrial ecosystems (Vitousek and Howarth, 1991) and
also most frequently limits agricultural productivity. In soils, the majority of N exists in
complex organic forms that must be mineralized (converted from organic N to NH4+ or
NO3-) prior to uptake by most agricultural
plants (Stevenson and Cole, 1999), although it
is increasingly recognized that most plants also
take up organic N (Jones et al, 2002; Schimel
and Bennett, 2004). Recent studies have demonstrated that the addition of biochar to surface mineral soils may directly influence N
transformations. Here we review the evidence
for the direct and indirect influences of biochar
on ammonification, nitrification, volatilization,
denitrification, N2O emission (see also Chapter
17) and N2-fixation, while providing potential
mechanisms that may be driving these
Ammonification and nitrification
Nitrogen mineralization is the process by
which organic N is converted to inorganic
forms (primarily NH4+ and NO3-). The conversion of organic-N to NH4+ is generically
termed ammonification. This process is driven
by a broad consortium of organisms capable of
enzymatic denaturation of proteins and the
removal of amide groups from organic compounds (e.g., amino acids and amino sugars).
Nitrification represents the oxidation of
organic N (via heterotrophic organisms) or
NH4+-N to NO3- by autotrophic bacteria and
Biochar C15.indd 427
Figure 15.4 The soluble PO4-3, NH4+ and
NO3- concentration in biochar produced from
Douglas-fir or Ponderosa pine wood or bark at
350 or 800ºC (from Gundale and DeLuca,
2006). Data meeting the assumptions of
normality were compared with one-way
ANOVA followed by the Student-NeumanKuels post hoc procedure where letters indicate
pairwise differences. Non-normal data were
compared using the Kruskal-Wallis (K-W)
archaea as well as certain fungi (Stevenson and
Cole, 1999; Leininger et al, 2006). Biochar
addition to temperate and boreal forest soils
has been found to increase net nitrification
11/3/2014 9:04:30 AM
rates in soils that otherwise demonstrate little
or no net nitrification (Berglund et al, 2004;
DeLuca et al, 2006); whereas, there has been
little evidence for such an effect in grassland
(DeLuca et al, 2006) or agricultural soils
(Lehmann et al, 2003; Rondon et al, 2007),
which may already accommodate an active
nitrifying community. Results from the literature are summarized in Table 15.1, which specifically focuses on ammonification and
nitrification resulting from biochar or activated
carbon amended soil samples, field plots, or
mesocosms, as compared to unamended
Several studies in forest ecosystems have
aimed to understand the mechanisms underlying increased nitrification following PCM
addition. Using forest soils with very low
inorganic N concentrations, DeLuca et al
(2002) showed that ammonification was
enhanced when a labile organic N substrate
(i.e. glycine) was added to the soil, indicating
that ammonification was substrate limited. In
contrast, the large increase in NH4+ that
resulted from the glycine addition did not
cause soil NO3- concentration to increase,
suggesting that nitrification was not substrate
limited (DeLuca et al, 2002); rather, nitrification rates exceeded ammonification rates. In
the same study, the injection of activated carbon into the organic horizon induced a slight
stimulation of nitrification (see Table 15.1),
but the injection of glycine with activated carbon consistently stimulated high rates of
nitrification, demonstrating that biochar in
some way alleviated the factor limiting nitrification (DeLuca et al, 2002; Berglund et al,
2004). Char collected from recently burned
forests (DeLuca et al, 2006; MacKenzie and
DeLuca, 2006) or generated in laboratories
under controlled conditions (Gundale and
DeLuca, 2006) were found to stimulate net
nitrification in laboratory incubations and in
short-term (24hr) nitrifier activity assays.
One possible mechanism is that activated car-
Biochar C15.indd 428
bon adsorbed organic compounds (and specifically terpenes) that either inhibited net
nitrification (White, 1991; Ward et al, 1997;
Paavolainen et al, 1998) or caused immobilization of NH4+ (McCarty and Bremner,
1986; Schimel et al, 1996; Ward et al, 1997;
Uusitalo et al, 2008). The rapid response of
the nitrifier community to biochar additions
in soils with low nitrification activity and the
lack of a stimulatory effect on actively nitrifying communities suggest that biochar may be
adsorbing inhibitory compounds in the soil
environment (Zackrisson et al, 1996) that
then allows nitrification to proceed. Similarly,
fire induces a short-term influence on N
availability, but biochar may act to maintain
that effect for years to decades after a fire. It
is also possible that the presence of biochar in
these forest soils enhances the numbers of
ammonia oxidizing bacteria by creating conditions conducive to their growth (increased
pH, reduced inhibitory compounds, microsites, redox potential, external electron transfer) (Ball et al, 2010).
In another study seeking to explain charinduced increased nitrification rates in nutrient poor conifer forests, DeLuca et al (2006)
evaluated gross nitrification rates in chartreated and untreated forest soils. Gross nitrification rates in the char-amended forest soils
were nearly four times that in the untreated
soil, demonstrating the stimulatory effect of
char on the nitrifying community rather than
reduced immobilization. We also found that
sterilized soils treated with biochar also exhibited a modest increase in nitrifier activity
(DeLuca, unpublished data) suggesting that
the oxide surfaces on biochar may stimulate
some quantity of auto-oxidation of NH4+
(DeLuca et al, 2006). Wood ash commonly
contains high concentrations of metal oxides
including CaO, MgO, Fe2O3, TiO2 and CrO
(Koukouzas et al, 2007). Exposure of biochar
to solubilized ash may result in the retention
of these potentially catalytic oxides on active
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Biochar C15.indd 429
11/3/2014 9:04:30 AM
Lab Biochar,
Ponderosa pine
Lab Biochar,
Ponderosa pine
(wood and
(wood and
Ponderosa pine,
Western MT
Scots Pine,
Scots Pine,
Ponderosa pine
Biochar Type
Ponderosa pine,
Western MT
Ponderosa pine,
Western MT
Glycine in lab,
incubation, 60
0.06 ± 0.02
2.8 ± 0.4
46 ± 6
(µg N g soil -1)
47 ± 4
(µg N g soil -1)
Glycine in lab,
14 days
20 ± 13
(µg N cap-1) ‡
40 ± 5
(µg N g soil -1)
(NH4)SO4 and
H2PO4 in lab,
15 days
Glycine in Field,
resin collected,
30 days
200 ± 100
1350 ± 50
5.5 ± 0.6
high 1.89
± 1.1
high 780 ± 302
dfb 16 ± 4
dfb 27 ± 3
dfw 11 ± 8
dfw 32 ± 8
low 0.12
± 0.03
ppb 20 ± 8
ppb 25 ± 6
low§ 410 ± 99
ppw 21 ± 4
70 ± 3
1200 ± 500
NO3- - N
ppw † 20 ± 5
700 ± 400
NH4+ - N
NH4+ - N
NO3- - N
Biochar Addition
150 ± 200
(µg N cap-1) ‡
Glycine in
resin collected,
30 days
Nutrient Source
and Incubation
Berglund et al
DeLuca et al
Gundale and
DeLuca et al
and DeLuca
Table 15.1 The effect of biochar (natural biochar, lab-generated biochar, or activated carbon) on nitrogen mineralization, nitrification and
immobilization from studies performed in soils of different ecosystems around the world
Biochar C15.indd 430
11/3/2014 9:04:30 AM
Biochar Type
Mineralizable N
None, Aerobic
49 days
Four feedstocks,
bark, Switch
grass, Digested
5 agronomic
soils (S, SiL),
30 to 70
(mg N kg-1)
26.8 ± 12.2
0.24 ± 0.08
30.7 ± 12.6
fresh0.63 ± 0.45
old 5.16 ± 3.50
1.74 ± 0.30
old 3.87 ± 0.76
Mineralizable N
1.05 ± 0.05
67 days ND
fresh 1.42 ± 0.21
0.6 ± 0.1
25 days 2.25
146 ± 42
NO3 - N
‡ Ionic resin analysis used approximately 1g mixed bed resin in nylon mesh capsules approximately 25.4mm in diameter
† ppw – Ponderosa pine wood; ppb – Ponderosa pine bark; dfw – Douglas-fir wood; dfb – Douglas-fir bark biochar produced at 350ºC
§ Low biochar application rate of 1,000kg ha-1, high application rate was 10,000kg ha-1
!! Sand (S) and silt loam (SiL) textured soils
15.0 ± 14.2
(mg N kg-1)
None, aerobic,
7 days
Perennial Rye
Eutric Camisol
Wales, UK
2.02 ± 0.31
3.70 ± 0.38
(mg N kg-1)
None, aerobic,
7 days
(mg L-1)
49 kg ha-1 UAN
aerobic, 25 days
Wheat Residue
20 ± 3
(µg N cap-1) ‡
0.20 ± 0.20
NH4 - N
NO3 - N
NH4 - N
Biochar Addition
Glycine in Field,
resin collected,
75 days
Nutrient Source
and Incubation
Winter Wheat
Kandsol Soil
Forested continued
Scots Pine,
Corn Production Pecan Shell
South Carolina
Table 15.1 continued
Novak et al
lower in most
Streubel et al
Not Reported Dempster et
al (2012)
Not Reported Dempster et
al (2012)
surfaces of the biochar (Le Leuch and
Bandosz, 2007). These oxide surfaces may in
turn effectively adsorb NH4+ or NH3 and
potentially catalyse the photo-oxidation of
NH4+ (Lee et al, 2005). More work is required
to understand these variable mechanisms on
nitrification in forest soils.
In contrast to forested ecosystems, biochar additions in agricultural systems have
yielded mixed results, partially based on the
variety of feedstocks tested in agricultural trials. Biochar additions to agricultural soils
have been found to reduce, have no effect or
in some cases increase net N mineralization
(Yoo and Kang, 2010; Streubel et al, 2011;
Güereña et al, 2013). Streubel et al (2011)
tested 5 different soils from the agronomic
regions of Washington state, ranging in texture from sand to silt loam (Table 15.1) and
amended with three rates of biochar (9.8,
19.5 and 39.0Mg ha−1), produced from four
different feedstocks (Douglas-fir wood and
bark, switchgrass stubble and digested fibre).
In this study, biochar had no significant
effect, or reduced N mineralization regardless of feedstock, for all but the sandy soil
amended at the highest quantity of switchgrass biochar, which had significantly higher
NH4+ and NO3- production. Contrasting
results were reported in alkaline, calcareous
agricultural soils amended at a rate of 1–10
per cent switchgrass biochar in a laboratory
incubation, wherein biochar applications
reduced NH4+ and NO3- accumulation
(Ducey et al, 2013). However, when soils are
treated with biochar produced from a N-rich
feedstock (e.g. swine manure), increases in
net N mineralization and net nitrification
have been observed in laboratory incubations
(Yoo and Kang, 2010). Importantly, barley
straw biochar amendments to the same soils
had no significant effect on N mineralization.
Using molecular analyses (TRFLP and 454
pyrosequencing) microbial response to biochar additions were studied in agricultural
Biochar C15.indd 431
soils; the presence of Nitrosovibro (NH4+ →
NO2-) was found to decrease in the presence
of biochar while Nitrobacter (NO2- → NO3-)
was observed to increase in the presence of
biochar (Anderson et al, 2012). However,
these shifts could have little consequence for
nitrification rates as molecular analyses have
also demonstrated little or no relationship
between gene abundance of ammonia oxidizing bacteria and rates of NO3- accumulation
(Ducey et al, 2013). Such results emphasize
the contrast between the strong positive
effects biochar amendment has on forest
soils, where little or no net nitrification occurs,
compared to absence of effects in agricultural
soils, that already exhibit inherently high
rates of net nitrification and NO3- accumulation (e.g. over 113mg NO3- -N kg-1 in the
control, Ducey et al, 2013) prior to biochar
additions. Interestingly, Nelissen et al (2012)
reported a significant increase in gross
ammonification and nitrification rates in
sandy soils amended with maize biochar with
the increase in nitrification being attributed
to greater substrate availability for autotrophic nitrifying bacteria.
The length of time that biochar resides in
the soil environment has also been shown to
affect N mineralization potential which may
be related to its occlusion with organic matter
over time as postulated by Zackrisson et al
(1996). Dempster et al (2012) found that
soils amended with ‘old’ biochar resulted in
greater inorganic N accumulation than soils
amended with ‘fresh’ biochar in different
agronomic soils from both Australia and the
UK (Table 15.1). This might have significant
implications for management practices that
are using biochar to retain inorganic N fertilizer on-site. Regular additions of ‘fresh’ biochar to agricultural systems might be needed
to help retain inorganic N fertilizers and this
practice may also sequester large amounts of
C. In contrast, Novak et al (2010) reported a
modest increase in net N mineralization when
11/3/2014 9:04:30 AM
fresh wood biochar was added to acidic agricultural soils. Biochar additions to agricultural soils of the tropics have been reported to
either reduce N availability (Lehmann et al,
2003) or to increase N uptake and export in
crops (Steiner et al, 2007; Cornelissen et al,
2012). Reduced N availability may be a result
of the high C:N of biochar and thus greater
potential for adsorption of NH4+ to biochar
which in turn reduces the potential for
N leaching losses and sustained higher N fertility over time in surface soils (Steiner
et al, 2007). Alternatively, a biochar-induced
increase in microbial biomass could result in
net immobilization of N as demonstrated in
N fertilizer trials on corn fields in New York
(Güereña et al, 2013).
Several studies have suggested that biochar can adsorb NH4+ from the soil solution
(Lehmann et al, 2011; Spokas et al, 2012;
Taghizadeh-Toosi et al, 2012a), thus reducing the availability of NH4+ for autotrophic
conversion to NO3-; however, this hypothesis
has not been adequately tested. Increased
sorption of NH4+ could, at least temporarily
create localized concentrations for microbial
use or plant uptake. Although reduced N2O
emissions with biochar additions have been
ascribed to increased sorption of NO3- by
biochar (van Zwieten et al, 2010), NO3- sorption to wood biochar has also been reported
to be insignificant (Jones et al, 2012). In addition to sorption of inorganic N, biochar also
has the potential to adsorb organic N compounds (amino acids, peptides, proteins)
which could reduce N net mineralization or
nitrification, or stimulate N-fixing microbial
growth around and within biochar particles,
coating them with N-retaining microbes.
Moreover, biochar in soil ultimately becomes
occluded with organic matter and may aggregate both mineral and organic matter fractions together into physically protected pools
(Brodowski et al, 2006), N present in these
organic pools would likely remain unavailable
Biochar C15.indd 432
for some period of time and be protected
against transformation.
Biochar from wood or other N-poor feedstocks is an N-depleted material with a high
C:N ratio; however, biochars generated from
N-rich feedstocks can serve as a N source
(Lehmann et al, 2006). It has generally been
found that some decomposition occurs when
fresh biochar is added to soil (Schneour,
1966; Spokas et al, 2009; Jones et al, 2011),
although it is also well known that wood biochar is highly recalcitrant (DeLuca and Aplet,
2008). It is therefore uncertain whether biochar amendments provide sufficient bioavailable C to stimulate immobilization
(although this argument is often cited).
Several studies have demonstrated increased
respiration rates with biochar additions to soil
and suggested that the biochar was being
degraded or that the biochar induced priming
of resident soil C or fresh C (Wardle et al,
2008; Spokas et al, 2009; Novak et al, 2010).
However, recent studies have shown that
increased CO2 evolution, which is usually
indicative of increased microbial activity,
upon biochar additions was actually partly
due to emission of inorganic C (i.e. carbonates) from within the applied biochar (Jones
et al, 2011). Low-temperature biochars, in
particular, would be more likely to induce net
nutrient immobilization, because they contain higher concentrations of bioavailable C
and residual bio-oils (Steiner et al, 2007;
Nelissen et al, 2012; Clough et al, 2013) or
surface functional groups (Liang et al, 2006)
that can serve as microbial substrates. Higher
temperature biochars, in contrast, contain
much higher concentrations of graphene
structures which are much more resistant to
microbial degradation and less residual volatiles. When biochars do provide a significant
concentration of bioavailable C (see Chapter
16), any immobilization stimulated by bio-
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char would likely lead to only short-term
reductions in available inorganic N pools that
could reduce nitrification and N2O emissions
(Chapter 17) (Steiner et al, 2007).
Gaseous nitrogen emissions
Over the past several years there has been an
increasing interest in understanding how biochar influences the gaseous soil N transformations in order to understand ecosystem N
budgets and effects of biochar management
on greenhouse gas emissions. Much interest
has focused on the influence of biochar on
N2O flux (i.e., it has a global warming effect
per molecule that is 298 times greater than
CO2) (Yanai et al, 2007; Spokas et al, 2009;
Clough et al, 2010; Cornelissen et al, 2012),
because of its importance as a greenhouse gas
(Hansen et al, 2005) and ozone-depleting
substance (Ravishankara et al, 2009). Several
studies have also addressed the influence of
biochar applications on denitrification and
NH3 volatilization potential to evaluate the
influence of biochar on N conservation in
agricultural soils (Jones et al, 2012;
Taghizadeh-Toosi et al, 2012a). Nitrous
oxide emissions from soil are associated with
the processes of nitrification and denitrification and this topic is covered in detail in
Chapter 17.
To date there have been a number of
papers devoted to assessing N2O generation
(as a greenhouse gas) on biochar-amended
agricultural soils, but far less effort geared
towards assessing the influence of biochar on
denitrification potential. Kammann et al
(2012) reported a significant decrease in N2O
emissions under ‘denitrification inducing
conditions’ following the addition of 50Mg
ha-1 of peanut hull biochar to an agricultural
Luvisol in Germany. However, when the biochar was added with 50kg N ha-1 as NH4NO3,
N2O emissions actually increased with biochar additions (Kammann et al, 2012). The
latter results are similar to the findings of
Biochar C15.indd 433
Jones et al (2012) which demonstrated an
increase in denitrification enzyme activity
DEA in north Wales where soils had been
treated with 50Mg wood biochar ha-1 and
100kg N ha-1 as NH4NO3 two years prior to
sampling. Studies have also demonstrated an
increase in denitrifying bacterial communities
with biochar additions to soil (Anderson et al,
2012; Ducey et al, 2013).
As indicated above, biochar applications
to agricultural or grassland soils that exhibit
high rates of nitrification have generally been
found to have no effect or a negative effect on
net nitrification rates (DeLuca et al, 2006;
Rondon et al, 2007; Dempster et al, 2012;
Jones et al, 2012) which would act as to
reduce denitrification potential. In contrast,
forest soils that otherwise express little or no
net nitrification respond to biochar or activated carbon additions by exhibiting an
increase in net nitrification and nitrifier activity (Berglund et al, 2004; DeLuca et al,
2006). Further, accumulation of NO3- in soils
treated with soluble C-rich biochar would act
to increase denitrification potential under
anaerobic conditions (McCarty and Bremner,
1992). Hydochar produced under hydrothermal conditions increases soluble C additions
and has the potential to increase respiration
(Kammann et al, 2012) and denitrification
potential. However, application of wood biochar likely results in a long-term net decrease
of soluble organic C (see Chapter 16), which
would in turn reduce denitrification
Several recent studies have also evaluated
the influence of biochar on NH3 volatilization
(Steiner et al, 2010; Doydora et al, 2011;
Jones et al, 2012; Taghizadeh-Toosi et al,
2012a, b; Chen et al, 2013). Ammonia volatilization in agricultural soils is favoured at
alkaline pH and when high concentrations of
NH4+ are present and is reduced in soils with
high CEC (Stevenson and Cole, 1999).
Biochar and biochar mixed with ash are
11/3/2014 9:04:30 AM
known to temporarily increase soil pH (Glaser
et al, 2002; Jones et al, 2012), but usually not
to a high enough level to increase NH3 volatilization. Taghizadeh-Toosi et al (2012a,b)
have shown instead that NH3 is effectively
sorbed to the surface of wood biochar, but
also demonstrate that it can be desorbed into
solution as NH4+ thereby preventing N losses
to the atmosphere.
Biochar additions to agricultural soils as
well as acid forest soils have been found to
reduce NH4+ concentrations (Le Leuch and
Bandosz, 2007; Taghizadeh-Toosi et al,
2012a) which reduces the potential for NH3
volatilization. Steiner et al (2010) found a
clear reduction in NH3 evolution during
poultry litter composting when biochar
amendment rates were 20 per cent (w/w).
Doydora et al (2011) found 50–60 per cent
reductions in NH4+ available for volatilization
when composting poultry litter was cut 1:1
with biochar prior to incorporation into soil.
This finding is supported to some degree by
Jones et al (2012) who found a clear capacity
of biochar to adsorb NH4+. Furthermore, in
field trials the researchers showed a reduction
in NH3 volatilization at rates of 50Mg biochar
ha-1, but not at 25Mg biochar ha-1 (Jones et al,
2012). In agricultural soils, it appears that
biochar generally results in a reduced presence of extractable NH4+, likely as a result of
sorption of soluble NH4+ to biochar surfaces.
However, the reduced occurrence of NO3formation in agricultural soils treated with
biochar suggests that the reduced oxidative
pathway may leave more NH4+ present and
available for immobilization. Further studies
are clearly required to determine the capacity
of biochar to reduce volatilization of soil or
applied NH4+.
Biological nitrogen fixation
Biological N2 fixation historically provided
the vast majority of N inflow into agroecosys-
Biochar C15.indd 434
tems (Galloway et al, 2008). Today it is mandatory in low-input agroecosystems where
external N inputs are minimal. A thorough
understanding of the influence of biochar on
both symbiotic and free-living N2 fixing
organisms is needed. The influence of biochar on N2 fixation in leguminous plants was
studied with some intensity over half a century ago, as researchers investigated the role
of charcoal as a soil amendment. Although
char was noted to have positive influence on
soil physical properties (Tryon, 1948), there
has been no consistent observed influence of
biochar on N2 fixing leguminous plants.
Table 15.2 provides a summary of results of
studies on the influence of PCM applications
on nodulation and N2 fixation in leguminous
plants. Several of these studies are described
in the following section.
Vantis and Bond (1950) found that the
addition of wood biochar to soils at a rate of 1
per cent (v/v) resulted in a reduction in the
number of nodules on clover, but increased
the total nodule mass and total N2 fixed in
Pisum sativum (L.). However, at higher rates
of biochar (greater than 2 per cent) there was
no effect or a negative effect of biochar on
nodulation (Vantis and Bond, 1950). Turner
(1955) found a significant increase in the
number of root nodules in clover (Trifolium
pratense L.), but only after an initial reduction
compared to the control. Their work went on
to show that boiled biochar further increased
nodulation and suggested that inhibitory compounds can be removed by various forms of
pretreatment of biochar (Turner, 1955) (this
treatment may have influenced phytohormone-like chemicals, see Chapter 14).
Investigation of composts with or without biochar added (5% w/w) as a growth medium
suggested that the biochar additions resulted
in a significant decrease in nodule number and
size (Devonald, 1982), however, there is no
discussion on pretreatment of the biochar or
its polyaromatic hydrocarbon (PAH) content.
11/3/2014 9:04:30 AM
Biochar C15.indd 435
11/3/2014 9:04:30 AM
Pisum sativum
Trifolium pratense
P. sativum
Medicago sativum
Phaseolus vulgaris
Glycine max
Trifolium repens
Trifolium repens
Wood biochar 2%
Wood biochar 4%
Wood biochar 8%
Activated carbon 1%
Animal biochar 2%
Wood biochar 1% –2% ¶
Wood biochar powder 1:1
Wood bark biochar ~ 1%
Wood biochar 3%
Wood biochar 6%
Wood biochar 9%
Chicken manure biochar ~ 0.4%
Chicken manure biochar ~ 0.8%
Wood biochar ~2.5%
Wood biochar ~ 5%
NA: Not available
NS: Not significant at P < 0.05
Response plant
Biochar type and rate
+ 39%
+ 25%
- 24%
NS or neutral
+ 8%
+ 45%
+ 37%
Growth response
- 39%
+ 97%
- 31%
- 11%
+ 25%
+ 350%
+ 250%
+ 64%
+ 42%
Quilliam et al (2012)
Quilliam et al (2012)
Tagoe et al (2008)
Rondon et al (2007)
Nishio and Okano (1991)
Devonald (1982)
Turner (1955)
Vantis and Bond (1950)
Table 15.2 Summary of research findings on the influence of biochar or activated C on growth, nodulation and N2 fixation in
leguminous crops. For each study, percentage change in individual variables were calculated relative to an experimental control. All
studies are pot trials with the exception of Quilliam which combines field application of biochar with a growth chamber pot trial
These findings are supported by more
recent studies that demonstrate a noted inhibitory effect of activated carbon on nodulation
in Lotus corniculatus (L.) (Wurst and van
Beersum, 2008). On the other hand, the
application of a nutrient-rich biochar (carbonized chicken manure) to silt loam soils in
a greenhouse experiment was found to
increase nodule number and mass in soybeans (Glycine max L.) and increase total N
yield (Tagoe et al, 2008). In a recent field
study (Quilliam et al, 2013), high rates of
wood biochar were applied to temperate agricultural soils, (total applications of 25, 50 and
100kg biochar ha-1), soils were collected and
placed in pots and seeded to clover (T.
repens). Clover plants grown in the presence
of high rates of biochar were observed to have
reduced numbers of nodules, but the mass of
nodules was increased as was nitrogenase
activity (Quilliam et al, 2013).
Rondon et al (2007) tested the effect of
adding different amounts of wood (eucalypt)
biochar to nodulating and non-nodulating
varieties of the common bean, Phaseolus vulgaris, inoculated with Rhizobium strains and
measured changes in N uptake using an isotope pool dilution technique. Biochar was
found to significantly increase N2 fixation and
bean productivity at application rates of 30 or
60g biochar kg-1 compared to a control, but
the highest application rate, 90g biochar kg-1
soil reduced bean productivity (Rondon et al,
2007). All biochar treatments increased the
percentage of N in bean tissue derived from
N2 fixation. The study further indicates that
biochar may stimulate N2 fixation as the result
of increased availability of trace metals such
as nickel (Ni), iron (Fe), boron (B), titanium
(Ti) and molybdenum (Mo). The highest
rates of biochar application decreased the
magnitude of the effect and if taken to the
extreme might interfere with N2 fixation.
It is possible that lack of consistent effects
of biochar on legume performance and nodu-
Biochar C15.indd 436
lation is due to differences in nutrient contents of the various biochars and their
respective potential to adsorb signalling compounds. There is certainly variation in the
nutrient content and sorption capacity of
various types of biochar. Nodule formation in
leguminous plants is initiated by the release of
signalling compounds, often flavonoids (Jain
and Nainawatee, 2002). Such polyphenolic
compounds are readily sorbed by biochar
(Gundale and DeLuca, 2006). This might
explain why some studies have shown that
activated carbon reduces nodulation, while
low sorption P-rich biochars increased nodulation, which is presumably the result of alleviating P-limitation of nodulating bacteria
with high P demands, such as Rhizobium spp.
(Rondon et al, 2007).
Numerous studies have been conducted
to evaluate the potential for increasing the
activity of free living N2 fixing bacteria in
agroecosystems, however, this has only been
directly evaluated in a few papers (Table
15.2). One methodological limitation to studying this is that biochar additions to soil likely
increase ethylene production (Spokas et al,
2010; see also Chapter 14), which is used as a
surrogate for N2 fixation activity in the acetylene reduction assay, the most commonly
used technique to estimate nitrogenase activity. It is well understood that excess soluble N
in the soil solution reduces N2 fixation rates in
free-living N2-fixing bacteria (Kitoh and
Shiomi, 1991; DeLuca et al, 1996) and available soil P can stimulate N2 fixation (Chapin
et al, 1991). Therefore, it is possible that the
activity of free living N2-fixing bacteria could
be increased by biochar-induced increases in
P solubility (Lehmann et al, 2003; Steiner et
al, 2007) and reduced soluble soil N concentrations (due to immobilization or surface
adsorption of NH4+). Biochar therefore
potentially represents a good carrier or
medium for the growth and proliferation of
free-living N2-fixing bacteria. Wood and
11/3/2014 9:04:30 AM
cellulose based biochars are low N media, yet
serve to sorb soil P (see Chapter 7).
Conversely, however, in low inorganic N
environments, biochar is known to stimulate
net nitrification (e.g., DeLuca et al, 2006)
and thus may ultimately down-regulate N2
fixation by free-living N2-fixing bacteria.
While many of these mechanisms may influence the activity of free living N2-fixing bacteria, very little work has been done on this
Biochar additions to soil have been found to
both increase and decrease the availability of
soil P (Steiner et al, 2007; Nelson et al, 2011).
After N, P tends to be the next major nutrient
limiting primary production in most ecosystems, except in the tropics where it is often
the primary limitation (subsequently limiting
microbial N2 fixation). Unlike N, there is little
evidence for the direct uptake of organic P by
plants and therefore soil organic matter containing organic P polymers (e.g. phospholipids, DNA, phosphorylated proteins etc.)
must be enzymatically broken down outside
the cell prior to the uptake of inorganic P (Pi).
Inorganic P is most commonly taken up by
plants in the HPO4-2 or H2PO4- form. Some
low molecular weight organic P can be
directly taken up by microbial cells (e.g.
adenosine phosphates), however, this pathway is probably small in comparison to the
uptake of Pi. In contrast to N, however, the
solubility and rate of diffusion of Pi in soils is
typically extremely low due to strong sorption
to the mineral phase (e.g. on Fe and Al oxyhydroxide surfaces) and its potential to form
mineral precipitates (e.g. Ca-P). Although
biochar itself can provide a readily available
source of P, it may also directly and indirectly
influence P behaviour in soil by a range of
other mechanisms including: (i) shifts in soil
pH; (ii) alteration in enzyme efficiencies; (iii)
Biochar C15.indd 437
formation of organo-mineral complexes that
increase P solubility; (iv) shifts in plant and
microbial community structure.
Release of P from biochar and impacts
on P leaching
Most biochars contain an appreciable amount
of P and the release of P from biochar has
long been recognized (Tryon, 1948). The P
contained in biochar feedstocks typically
ranges from 0.01–0.1 per cent for wood residues, 0.1–0.4 per cent for crop residues and
0.5–5 per cent for manures and biosolids
(Barrelet et al, 2006; Liu et al, 2011; Wang et
al, 2012). Within each feedstock, P can be
contained in different chemical forms which
may affect its subsequent availability after
entering the soil. When plant residues are
pyrolysed, organic C begins to volatilize at
approximately 100°C; whereas, P does not
volatilize until approximately 700°C (Knoepp
et al, 2005). Combustion or charring of
organic materials can greatly enhance P availability from plant tissue by disproportionately
volatilizing C and by cleaving organic P
bonds, resulting in a residue of soluble P salts
associated with the charred material. Pyrolysis
leads to the formation of a range of mineral P
forms of which complexes with Fe, Al, Ca
and Mg predominate. It is likely that biochars
derived from biosolids (sewage sludge) will
have a high proportion of Fe and Al phosphates which will be less soluble than Ca and
Mg phosphates. Biochar therefore contains
three pools of P, one which is freely soluble,
one which is strongly bound to Fe and Al and
one which remains organically bound as a
residue of the original feedstock. The proportion of each is dependent on feedstock and
pyrolysis conditions (covered in detail in
Chapter 6) (Liu et al, 2011), however, a
number of studies have shown that a large
proportion of the P contained in woodderived biochars is immediately soluble and
readily released into soil solution (e.g.
11/3/2014 9:04:30 AM
Gundale and DeLuca, 2006). Experimental
biochar addition rates of 10–50kg biochar
ha-1 would result in the addition of 1 to 2500
kg P ha-1 depending on the feedstock of which
approximately 5–20 per cent can be assumed
to be available depending on feedstock (see
Chapter 7). When considering that typical
synthetic fertilizers are added at 5 to 50kg P
ha-1 there is considerable potential for overapplication with some biochars. However,
there has been little evidence for increased
leaching of P in laboratory trials (Laird et al,
2010; Borchard et al, 2012a; Schultz and
Glaser, 2012). Quilliam et al (2012) described
a significant increase in available soil P (0.5M
acetic acid extraction) with reapplication of
wood biochar (25 and 50Mg biochar ha-1) on
top of an existing biochar application of 25
and 50Mg biochar ha-1 (resulting in a 0, 25,
50, 25+25 and 50+50Mg biochar ha-1 experiment) in an agricultural Inceptisol. However,
this noted increase in P availability did not
translate to an increase in foliar P concentrations in the test crop (Quilliam et al, 2012).
Effect of biochar on phosphatase enzymes
and P solubilizing bacteria
Biochar can induce significant shifts in the
size, structure and activity of the soil microbial community (Lehmann et al, 2011).
Despite the significant amount of functional
redundancy in the microbial population, this
may cause changes in rates of P cycling.
Biochar has been observed to influence mycorrhizal colonization of plant roots which in
turn may alter soil P uptake (Warnock et al,
2007; Lehmann et al, 2011); however, this
topic is covered in detail in Chapter 13.
Short-term laboratory studies have
shown that biochar addition induces an
increase in phosphatase activity (Bailey et al,
2010; Yoo and Kang, 2010; Jindo et al, 2012)
which would increase the release of P from
soil organic matter and organic residues.
Other studies, however, have revealed only
Biochar C15.indd 438
minor influence on soil phosphatase (Jones
et al, 2010). Molecular analyses of soils
amended with biochar revealed increased
presence of bacteria genes for genera of bacteria that have been observed to produce P
solubilizing compounds (Hamdali et al, 2008;
Anderson et al, 2012); however, this is indirect evidence for biochar induced enhancement of P solubilizing bacteria and further
studies are needed to determine if there is any
direct influence on bacterial production of P
solubilizing compounds.
In addition to directly releasing soluble P,
biochar can have a high CEC (Liang et al,
2006). Sorbed Al, Fe or Ca may influence P
solubility (see page 439 and Chapter 9). It
has been demonstrated that fresh biochar has
an abundance of transient anion exchange
capacity in the acid pH range (Cheng et al,
2008), which can be in excess of the total
CEC of the biochar. These exchange sites
have the potential to compete with Al and Fe
oxides (e.g. gibbsite and goethite) for sorption of soluble P, similar to that observed for
humic and fulvic acid extracts (Sibanda and
Young, 1986; Hunt et al, 2007). To date,
there is a noted lack of studies evaluating the
effect of biochar and variable charge surfaces
(short-term anion exchange capacity) on P
cycling and availability.
As biochar ages, the positive exchange
sites on biochar surfaces decline and negative
charge sites develop (Cheng et al, 2008). The
biochemical basis for the high CEC is not
fully understood (see Chapters 5, 6 and 9),
but is likely due to the presence of oxidized
functional groups (such as carboxyl groups),
whose presence is indicated by high surface
O/C ratios that form on the surface of charred
materials following microbial degradation
(Liang et al, 2006; Cheng et al, 2008; Preston
and Schmidt, 2006). Phosphorus availability
and recycling may be influenced by the biochar CEC over long time scales and in soils
that have inherently low exchange capacities.
11/3/2014 9:04:30 AM
By reducing the presence of free Al+3 and
Fe+3 near root surfaces, biochar may promote
the formation and recycling of labile P fractions. This is also an area of research that
deserves greater attention.
A significant component of the P cycle consists of a series of precipitation reactions that
influence the solubility of P, ultimately influencing the quantity of P that is available for
uptake and actively recycled between plants
and microbes. The degree to which these precipitation reactions occur is strongly influenced by soil pH, due to the pH dependent
activities of the ions responsible for precipitation (Al3+, Fe3+ and Ca2+) (Stevenson and
Cole, 1999). In alkaline soils, P solubility is
primarily regulated by its interaction with
Ca2+, where a cascading apatite mineral pathway develops. In acid soils, P availability is
primarily regulated by its interaction with Al3+
and Fe3+ ions, where highly insoluble Al- and
Fe- phosphates form. Biochar may influence
precipitation of P into these insoluble pools
by altering the pH and thus the strength of
ionic P interactions with Al3+, Fe3+ and Ca2+
(Lehmann et al, 2003; Topoliantz et al, 2005)
or by sorbing organic molecules that act as
chelates of metal ions that otherwise precipitate P (DeLuca, unpublished data, Figure
15.4). The latter pathway may be of particular importance for plant P availability if biochar enhances formation of stabilized organic
matter from decomposing organic matter as
suggested by some composting studies
involving biochar (Dias et al, 2010) (see also
Chapter 25), or by altered SOC turnover patterns in Terra Preta (Liang et al, 2010).
Numerous studies have demonstrated
that biochar can modify soil pH, normally by
increasing pH in acidic soils (Mbagwu, 1989;
Matsubara et al, 2002; Lehmann et al, 2003).
There are few, if any, studies that have demonstrated a reduction in pH with biochar
Biochar C15.indd 439
addition in alkaline soils, however, the addition of acid biochar to acidic soils has been
observed to reduce soil pH (Cheng et al,
2006). An increase in pH associated with
adding biochar to acid soils are due to an
increased concentration of alkaline metal
(Ca2+, Mg2+ and K+) oxides in the biochar
and a reduced concentration of soluble soil
Al3+ (Steiner et al, 2007). Adding these alkaline metals, both as soluble salts and associated with biochar exchange sites, is likely the
single most significant effect of biochar on P
solubility in the short term, particularly in
acidic soils where subtle changes in pH can
result in substantially reduced P precipitation
with Al3+ and Fe3+. In contrast, adding biochar (and associated ash residue) to neutral
or alkaline soils may have a limited effect on P
availability because adding alkaline metals
would only exacerbate Ca driven P
In addition to its effect on soil pH, biochar may also influence the bioavailability of
P through several other mechanisms associated with P precipitation, such as biocharinduced surface sorption of chelating organic
molecules. Biochar is an exceptionally good
surface for sorbing polar or non-polar organic
molecules across a wide range of molecular
mass (Sudhakar and Dikshit, 1999; Schmidt
and Noack, 2000; Preston and Schmidt,
2006; Bornermann et al, 2007). Organic molecules involved in chelation of Al3+, Fe3+ and
Ca2+ ions will potentially be sorbed to hydrophobic or charged biochar surfaces so that, in
the long run, organo-biochar or organo-mineral-biochar complexes begin to form over
time that may aid in the retention and
exchange of soluble P around aged biochar
particles (Briones, 2012; Joseph et al, 2013).
However, in the short term, decreased P solubility may result via the following mechanisms. Examples of such chelates include
simple organic acids, phenolic acids, amino
acids and complex proteins or carbohydrates
11/3/2014 9:04:30 AM
(Stevenson and Cole, 1999). The sorption of
chelates may have a positive or negative influence on P solubility. A clear example of this
type of interaction is provided in Figure 15.5.
Here, two compounds that have been
reported as possible allelopathic compounds
released as root exudates from Centaurea species: catechin and 8-hydroxy-quinoline
(Vivanco et al, 2004; Callaway and Vivanco,
2007) have also been reported to function as
potent metal chelates (Stevenson and Cole,
1999; Shen et al, 2001) that may indirectly
increase P solubility. Catechin effectively
increased P solubility in an alkaline (pH 8.0)
calcareous soil and the 8-hydroxy-quinoline
increased P solubility when added to an acidic
(pH 5.0) Al-rich soil (Figure 15.5). The
addition of biochar to these soils eliminated
the presence of soluble chelate in the soil system and in turn eliminated the effect of the
chelate on P solubility. This interaction may
explain the observed reduction in P sorption
by ionic resins with increasing biochar application rates in the presence of actively growing Koleria macrantha (Gundale and DeLuca,
2007). Such indirect effects of biochar on P
solubility would vary with soil type and vegetative cover and underscores the complexity
of plant-soil interactions.
To date, there have been few studies that
have focused on the influence of biochar soil
amendments on soil S transformations. Sulfur
plays an extremely important role in the biochemistry of soils and the physiology of plants
(Stevenson and Cole, 1999; Marschner,
2006). Sulfur is a component of two amino
acids (cysteine and methionine), is required
in protein synthesis and is a fundamental
component of energy transformations in all
living organisms. Sulfur also represents a
source of energy for autotrophic organisms
and an alternative electron acceptor for oxi-
Biochar C15.indd 440
dative decomposition under anaerobic conditions (Stevenson and Cole, 1999). Although
biochar produced from high S feedstocks has
the potential to release S into the soil solution
(Uchimiya et al, 2010), there is little evidence
for enhanced oxidation or reduction of soil S
with biochar applications. Given the similarities between the S and N cycles (Stevenson
and Cole, 1999), there is certainly the potential for biochar to influence S mineralization
and oxidation activity in the soil. Although
the majority of soil S originates from the geologic parent material, most soil S exists in an
organic state and must be mineralized (converted from organic S to SO4-2) prior to plant
uptake (Stevenson and Cole, 1999). Organic
S exists as either ester sulfate or as carbonbonded S, the latter having to be oxidized to
SO4-2 prior to plant uptake. To date, no studies have directly assessed the influence of biochar on transformations of organic S already
in agricultural or forest soils (separately from
S release from the biochar discussed below).
However, numerous studies involving biochar or biochar additions to soils have
recorded changes in the soil environment that
suggest that biochar additions could change
the turnover of soil organic S.
One of the few studies that have directly
investigated the influence of biochar on S
availability was conducted with two soil types,
four crop residue amendments and performed
in the laboratory in PVC columns (Churka
Blum et al, 2013). In this study, S, C and N
mineralization was assessed following addition
of corn husk biochar to soil compared with
fresh residues of corn husks, pea and rape residues. Although C mineralization and N mineralization were notably low with the biochar
amendment, the highest rate of S mineralization for all amendments was observed with the
corn husk biochar. The authors conclude that
the release of S from the residues is likely a
function of the S compounds within the residues. The high concentrations of ester S and
11/3/2014 9:04:30 AM
Figure 15.5 Soluble P leached from columns filled with (a) calcareous soil (pH = 8) amended with
catechin alone or with biochar or (b) acid Al rich soil (pH = 6) amended with 8-hydroxy quinoline
alone or with biochar (DeLuca, unpublished data). Studies were conducted by placing 30g of soil
amended with 50mg P kg-1 soil as rock phosphate into replicated 50mL leaching tubes (n = 3). Soils
were then treated with nothing (control), chelate, or chelate plus biochar (1% w/w), allowed to incubate
for 16 h moist and then leached with 3 successive rinsings of 0.01M CaCl2. Leachates were then
analysed for orthophosphate on a segmented flow Auto Analyzer III. Data were subject to ANOVA by
using SPSS
Biochar C15.indd 441
11/3/2014 9:04:30 AM
soluble SO4-2 in the biochar amendment combined with the modest increase in biotic activity with the biochar amendment suggests that
soluble SO3-2 and SO4-2 readily liberated from
the ester S allowed for rapid accumulation of
inorganic S in soils treated with biochar
(Churka Blum et al, 2013).
Biochar additions to acid agricultural
soils have been observed to yield a net
increase in soil pH (see Chapter 7), potentially as a function of the alkaline oxides
applied along with the biochar or potentially
as a result of the influence on free Al:Ca
ratios in soils amended with biochar (Glaser
et al, 2002; Topoliantz et al, 2005). Sulfur
mineralization is favoured at slightly acid to
neutral pH. Sulfur mineralization rates have
been found to increase following fire in pine
forest ecosystems (Binkley et al, 1992),
much the same as that observed for N
(Smithwick et al, 2005). Separating the
effect of fire from the effect of the natural
addition of char is difficult, but this effect is
most likely due to the release of soluble S
from litter following partial combustion during fire or heating events at temperatures in
excess of 200ºC (Gray and Dighton, 2006).
Sulfur oxidation is carried out by both
autotrophic (e.g. Thiobacillus spp.) and heterotrophic organisms. Sulfur oxidation by
acidophilic Thiobacillus spp. would not be
favoured by pH increases induced by the
presence of biochar. However, these autotrophic organisms have uniquely high
requirements for certain trace elements that
are in relatively high concentrations in biochar (see Chapter 13) and are increased in
soil when biochar is added (Rondon et al,
2007). Biochar additions to soil that ultimately reduce the surface albedo of mineral
soils and result in faster warming of soils in
springtime (Meyer et al, 2012) may in turn
increase S oxidation or mineralization rates
(Stevenson and Cole, 1999).
Biochar C15.indd 442
The utility of biochar, activated carbon
and graphite in the high temperature,
chemical reduction of SO2 to C adsorbed
episulfide has been studied intensely
(Humeres et al, 2005) to evaluate these
materials as scrubbers for removing SO2
from pollution streams. This has limited
relevance for soils, except that the very salts
demonstrated to enhance catalytic reduction of SO2 in the presence of activated carbon (e.g. Na+ and NO3-) are often present
in soils in relatively high concentrations.
However, there are no published reports of
enhanced SO4-2 reduction in soils in the
presence of biochar.
Biochar additions to mineral soils may
also directly or indirectly affect S sorption
reactions and S reduction. However, as
with NO3-, non-aged, production-fresh biochar may lack any significant capacity to
adsorb SO4-2 (Borchard et al, 2012b). As
noted in Chapter 5, biochar improves soil
physical properties through increased specific surface area, increased water holding
capacity and improved surface drainage.
Increased soil aeration through these
improvements in soil physical condition
would in turn reduce the potential for dissimilatory S reduction (Stevenson and
Cole, 1999). Sulfur is readily adsorbed to
mineral surfaces in the soil environment
and particularly to exposed Fe and Al
oxides. Once Fe and Al have been sorbed to
biochar surfaces, SO4-2 may interact with
the exposed metal oxides. Conversely,
organic matter additions to soil have been
shown to reduce the extent of SO4-2 sorption in acid forest soils (Johnson, 1984),
therefore biochar amendments could act to
increase solution concentrations of S in
acidic iron-rich soils. The lack of studies
devoted to the evaluation of S transformations following biochar addition to soils
calls for additional research in this area.
11/3/2014 9:04:30 AM
Future research directions
The application of biochar to agricultural
soils has the potential to greatly improve soil
physical, chemical and biological conditions.
In this chapter we reviewed biochar as a modifier of soil nutrient transformations and discussed the known and potential mechanisms
that drive these modifications. Biochar additions to soils may directly or indirectly alter
nutrient cycling. Biochar applications may
increase NH4+ retention in soil and have often
been observed to increase N uptake by crop
plants; however, there is little evidence for an
increase in N availability following crop harvest. It is important to note that biochar has
few negative impacts on soil nutrient cycling.
The observed increases in net nitrification in
forest soils occur at a level that would have
minimal influence on net N leaching and
N2O emissions. Although biochar additions
could increase net ammonification observations have not been consistent. While P solubility appears to generally increase with
biochar additions, this may be primarily a
result of direct P addition with the applied
biochar than altered P cycling. There is a distinct need for studies directed at explaining
mechanisms for increased P uptake with biochar additions to agricultural soils. It is possible that biochar additions to soils stimulate
mycorrhizal colonization, which may increase
P uptake, but when applied with P-rich materials, this effect may be lost. There is a great
need for additional studies that elucidate the
effect of biochar on soil nutrient transformations, in the short term directly after application as well as in the long term, e.g. in aged
PyC-rich soils. Some key areas that require
attention include: (1) Under what conditions
does biochar stimulate or reduce N mineralization, nitrification and immobilization in different ecosystems? (2) Does NH4+ adsorption
by biochar greatly reduce N availability or
does it concentrate N for plant and microbial
use? (3) By what mechanisms does biochar
addition to mineral soils stimulate P availability? (4) Do enzymes sorb to biochar and
retain their activity? (5) How does biochar
affect S availability and by what
mechanism(s)? The answers to these questions can only be obtained through rigorous
investigation of biochar as a soil conditioner
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