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FRONTIERS IN SOILS AND SEDIMENTS •REVIEW ARTICLE
The properties and functions of biochars in forest ecosystems
Yu Luo
1
&Zhuyun Yu
1
&Kaile Zhang
1
&Jianming Xu
1
&Philip C. Brookes
1
Received: 7 April 2016 /Accepted: 12 June 2016 /Published online: 25 June 2016
#Springer-Verlag Berlin Heidelberg 2016
Abstract
Purpose The production of large quantities of biochar from
natural fires has been a part of human history for millennia,
causing CO
2
emissions to the atmosphere and exerting long-
term effects on soil processes. Despite its potential importance
and recent work reflecting the wide interest in biochar, a gen-
eral review of our deep understanding of biochar functions
within forest soils is currently lacking. Gaps in research
knowledge in this field are identified in this paper.
Materials and methods This paper summarizes recent re-
search to provide a better understanding of the concentrations,
distribution, and characteristics of biochar produced from for-
est wildfire and its influences on soil processes. Perspectives
and recommendations for future research on biochar in post-
fire forest soils are also discussed.
Results and discussion The concentration, distribution, and
characteristics of biochar produced from forest wildfire large-
ly depend on forest landscapes, regional climates, and mostly
its feedstock and fire history, like, its duration and severity.
The influences of biochar on soil processes, particularly car-
bon and nitrogen transformations and cycling, like, nitrifica-
tion and nitrous oxide emissions reduction (Clough and
Condron, J Environ Qual 39:1218–1223, 2010), are also de-
termined mainly by the fire temperature and raw materials.
Mechanisms can be attributed to the adsorption of organic
compounds and nutrients or changed microenvironment,
termed as charsphere, by biochar. We also identify the micro-
bial mechanisms involved in the biochar-containing soils.
Keywords Biochar .Carbon sequestration .Distribution .
Forest .Microbial mechanisms .Nitrogen .Wildfire
1 Introduction
Fire has profoundly modified 40 % of the Earth’slandsurface
historically and is considered to be a natural phenomenon in
many forests (Bowman et al. 2009; David et al. 2009).
Currently, fire occurs in approximately 345 to 464 Mha per
year (Randerson et al. 2012). Fire often causes vast damage to
soil, e.g., damaging soil structure and porosity, leading to
losses of nutrients through volatilization, meanwhile, emitting
up to from 1.9 to 2.5 Pg carbon (C) to the atmosphere every
year (Santin et al. 2015a,b). This C loss during the fire will
contribute to global warming. However, in the longer term,
regenerating vegetation after the fire will compensate for the
loss of C caused by fire; thus, wildfire can be considered as a
Bnet zero C emission event^(Santin et al. 2015a,b).
Meanwhile, fire produces large amounts of biochar, between
the range of 116–385 Tg C each year (Santin et al. 2015a,b),
derived from roughly 1 % of the total world above-ground
biomass in forest system being incompletely combusted an-
nually (Ohlson et al. 2009).
Biochar is usually defined as a carbon-rich product pro-
duced by thermal decomposition of plant or animal residues
under zero or limited oxygen supply at temperatures above
300 °C (Lehmann 2007). Here, we use the term biochar to
Responsible editor: Zhihong Xu
Electronic supplementary material The online version of this article
(doi:10.1007/s11368-016-1483-5) contains supplementary material,
which is available to authorized users.
*Jianming Xu
jmxu@zju.edu.cn
1
Institute of Soil and Water Resources and Environmental Science,
College of Environmental and Resource Sciences, Zhejiang
Provincial Key Laboratory of Agricultural Resources and
Environment, Zhejiang University, Hangzhou 310058, China
J Soils Sediments (2016) 16:2005–2020
DOI 10.1007/s11368-016-1483-5
mainly refer to the carbonized plant material produced during
natural fire events. In a few cases, we also refer to the solid by-
product produced during the industrial pyrolysis of products
derived from plant and animal sources. Production of biochar
as charcoal has been a part of human history for millennia.
While fire causes CO
2
emissions to the atmosphere, the bio-
char produced after fire can be represented as negative fluxes
if considering a whole fire cycle, including fire-caused emis-
sion, recovery of vegetation, and biochar C remaining in soil,
from a few years to many decades (Santin et al. 2015a,b).
Compared with agricultural land, recent studies relevant to
biochar pay little attention to the forest ecosystems (Mitchell
et al. 2015). Biochar effects in agricultural soils are different
from the forest soils, as firstly, the soils in forest contain much
higher organic C and more diversity in microbial community;
secondly, the properties of biochar produced by forest fire are
also more heterogeneous than the industrial-produced biochar
following the standard protocol. Thirdly, most of agricultural
land received biochar only recently, while forest biochar
exited before human history; the biochar effects on the forest
ecosystems would be a much longer timescales than agricul-
tural land. The last but not the least, the biochar produced in
forest systems usually is a mixture product with the combina-
tion of plant and soil; instead, agricultural land received bio-
char after its production and not affected by fire. Thus, we
shouldbearinmindthatfire-derivedimpactsandfire-
derived biochar impacts are different; in most of cases, fire-
derived impacts are the short term and fire-derived biochar
impacts are relatively long term considering the recalcitrance.
However, it is difficult to distinguish these two effects on
forest soil system; thus, in this review, most of biochar effects,
if not specify, do not exclude the fire-derived impacts.
Despite its potential importance and recent work covering
diverse aspects reflecting wide interest in biochar (Sohi et al.
2010), a general review of our deep understanding of biochar
functions within forest soils is currently lacking. Our goal is to
summarize recent research to provide a better understanding
of the distribution, characteristics, and properties of biochars
produced from forest wildfire; their effect on soil processes;
and the microbial mechanisms involved. Finally, we provide
perspectives and recommendations for future research.
2 Distribution and characteristics
Recent research has investigated the spatial distributions and
properties of biochar produced from forest fires, combined
with analysis of their physical and chemical properties
(Kurth et al. 2006; Kane et al. 2010; Ohlson et al. 2013).
This has provided valuable information about its measurement
and effects, toward a better understanding of the distribution
and mechanisms involved in soil physio-chemical and
biological process in soils containing biochar produced from
wildfire.
2.1 Spatial distribution
Fire affects forest ecosystems significantly over a wide
range of spatial and temporal scales. Biochar spatial distri-
bution with specific patterns along the fire reflects clear
distinctions between fire-prone and fire-resistant regions.
Ohlson and Tryterud (2000) found that nearly half of
fire-produced biochars were less than 1 m within the fire
edge in their experiment. Similarly, Lynch and Clark
(2004) found that less than 1 % of biochar was moved
further than 20 m from the fire edge in the field by, for
example, water or wind. Biochar contents in soils also vary
significantly across different types of forest landscapes and
regional climates (Ohlson et al. 2009; Kasin and Ohlson
2013). A geo-statistical approach to present a detailed anal-
ysis of biochar contents in a boreal forest estimated a mean
value of 179 g C m
−2
(Ohlson et al. 2013). Kane et al.
(2010) also investigated soil organic carbon (SOC) and
biochar contents in black spruce forests in different land-
scapes in Alaska, USA, and he showed that the mean
biochar stocks in surface mineral soils (340 g C m
−2
)were
higher than in organic soils (170 g C m
−2
)inborealfor-
ests. The biochar pool in soil across large scales was wide-
ly investigated (Balshi et al. 2009; Ohlson et al. 2009);
meanwhile, the level of variation over small spatial scales
within a given forest landscape has been less studied.
Ohlson et al. (2013) calculated that biochar contents in a
boreal forest ranged from 34 to 1646 g C m
−2
within a
distance of 10 cm in soil, reflecting the great variability in
the size of the biochar pool over small spatial scales.
Eckmeier et al. (2007) did an experimental burning to sim-
ulate Neolithic slash and burn in deciduous forests and
found that the spatial distribution of biochar was highly
variable, with a range of 15–90 % of variation within
square meter. This estimation uncertainty due to the high
spatial variability makes it difficult to include biochar as
one of the sinks of atmospheric CO
2
in the IPCC guide-
lines (Turcios et al. 2016).
It was generally believed that most biochar is present in the
upper soil layers, mainly occurring in the surface between the
soil humic and mineral layers. This might be because biochar
is little moved into deep layer. It was reported that the content
of biochar C was close to the detection limit in the parent loess
and subsoil, suggesting that there was little if any vertical
translocation (Major et al. 2010). Only 1 % of biochar moved
downward in humid tropical soils (Major et al. 2010).
However, Turcios et al. (2016) found that up to 60 % of the
total biochar was stored in the top 30-cm depth, while 40 % of
biochar was stored below 30 cm, suggesting that newly
formed biochar can move into soil profile. Biochar was
2006 J Soils Sediments (2016) 16:2005–2020
observed to be dispersed along the soil vertical profile under
forests (Fig. 1). The depth of biochar accumulation depends
upon when fire occurred. Usually, the longer fire on forest
causes biochar accumulation at deeper depths (Fig. 1b), while
the young biochar produced from humified SOC and litter
after fire accumulates in the top humic layer (Fig. 1a). The
downward movement of biochar can be caused by bioturba-
tion; this can protect biochar from oxidation in future fires and
from dispersion by erosion. In contrast, Czimczik and
Masiello (2007) found that a fire-prone forest soil apparently
had no ability to store biochar, attributed to the lack of
burrowing organisms, like, earthworms, which incorporate
biochar into the mineral soil and protect it from oxidation by
the next fire. In contrast to preservation, the downward move-
ment of biochar in soil could also explain the rapid loss and
short mean residence time (<100 years) detected in the upper
layer soil (Zimmermann et al. 2012; Santin et al. 2015a,b).
Also, biochar was reported to be lost from soils via erosion
(Dittmar and Paeng 2009;Guggenbergeretal.2008;Rumpel
et al. 2006), suggesting that biochar mobility in dissolved and
colloidal phases is an important pathway to the ocean
sediments. Also, Jaffé et al. (2013) suggest that the mobiliza-
tion of biochar contributes to the dissolved organic carbon
(DOC) in ocean, accounting for about 10 % of the global
riverine flux of DOC.
2.2 Integration of biochar into soil particles
Biochar particles exhibit different morphologies ranging
from spherical to irregular shapes and from smooth to
rough surfaces, and particle sizes decline with time.
Zackrisson et al. (1996) sampled over 12 sites in northern
Sweden, and they found that the proportion of particle sizes
>1.6 mm declined and the proportion of particles <0.5 and
0.5–1.6 mm increased over 350 years. In another study, two
thirds of the biochar particles were larger than 2-mm diam-
eter in temperate deciduous forests (Eckmeier et al. 2007).
Biochar particles with similar morphologies were found in
different soil fractions, indicating that biochar exists in all
particle size fractions of soils within soil mineral complexes.
It was reported that biochar mainly occurred in a mineral-
associated form, which was highly dispersed in organo-
mineral forest soils (Filimonova et al. 2014). Other work
also showed that biochar was in the free light fraction and
20 % in aggregate-occluded and only 6 % in mineral-
associated fractions (Singh et al. 2014). However,
Brodowski et al. (2006) considered that instead of existing
in the light fraction, biochar mainly occurred in
microaggregates, and this was assumed to be one main
mechanism providing biochar stability. Within organic–min-
eral soil interfaces, biochar was found mostly with Fe or Al
oxides and clay minerals (Brodowski et al. 2005). We sam-
pled soil in the largest forest in China (Daxinanling) after
different fire periods and found that biochar distributed
mainly between plant roots and root-associated organic mat-
ter (Fig. 2). Knowledge about biochar distribution in com-
plex organo-mineral soils and different soil fractions remains
sparse; especially, how fire history affects biochar distribu-
tion in soil fractions needs further research, so that we can
better understand the recalcitrant properties of biochar
(Santos et al. 2012).
2.3 Factors determining biochar properties
Information about past burning can be obtained, mainly be-
cause biochar properties and distribution are determined by
the fire history and can reflect the extent and frequency of
fires and the plant biomass and species involved (Glaser
et al. 1998,2002; Keech et al. 2005). For example, Wolf
et al. (2013) measured the degree to which the chemical
signatures provided a fingerprint to predict fire temperature
using biochar properties including benzene polycarboxylic
acids (BPCAs), organic C content, and nitrogen content.
Thus, burning and raw materials during fire are the two
major factors that determine biochar properties. For exam-
ple, fire combustion will volatilize organic N, S, and O and
concentrate other elements (e.g., Na, P, and K) and organic
molecules; the extent of loss and accumulation largely de-
pends upon temperature. In general, at temperatures above
200 °C, biochar contains fewer O-alkyl structures, such as
cellulose, and more aromatic C, while at a charring temper-
ature of 350 °C or above, polyaromatic compounds are
Fig. 1 Vertical accumulation of
biochar with 40-year (a)and2-
year (b) fire history. Taken by the
authors in Daxinganling forest,
China
J Soils Sediments (2016) 16:2005–2020 2007
formed (Baldock and Smernik 2002). Fire is even consid-
ered as the first principal parameter mediating biochar prop-
erties, compared to combustible material; for example, in the
Ponderosapine(Pinus ponderosa)–Douglas–fir
(Pseudotsuga menziesii) ecosystem, both species and differ-
ent components (wood or bark) were of minor importance
compared to differences in biochar properties caused by dif-
ferent temperatures related to fire severity during fire events
(Gundale and DeLuca 2006).
Other factors affecting biochar properties include dura-
tion of charring (Braadbaart and Poole 2008), fuel (Wolf
et al. 2013), and even the wind speed, which affects O
2
supply and moisture (Braadbaart and Poole 2008). This
raises questions concerning studies under laboratory con-
ditions, where oxygen availability, maximum and mean
temperatures, duration of exposure, and soil moisture con-
tents are normally carefully controlled. This stability of
these parameters is in contrast to a natural fire event dur-
ing which they are highly variable in space and time
(Certini 2005). Therefore, such fire-induced transfers of
organic matter among forest ecosystems are not reflected
in laboratory experiments (Alexis et al. 2010). Thus, re-
search into the characteristics of biochar in in situ forest
ecosystems is valued even though they are still limited
(Knicker et al. 2008).
2.4 Methods adopted for biochar quantification
Quantification of biochar in soil compartments is crucial for
determining the mean residence time of pyrogenic C in the
environment, which raises another main concern about the
methods adopted to quantify biochar in soils and sediments.
Various approaches have been used (Electronic
Supplementary Material, Table S1), including hypespectral
imaging (Tong et al. 2015) and molecular marker methods,
solid-state
13
C nuclear magnetic resonance (NMR) spectros-
copy (Soucémarianadin et al. 2015),
13
CCPMASNMRspec-
troscopy (Rumpel et al. 2007), and
129
Xe NMR spectroscopy
(Filimonova et al. 2014). However, the results obtained from
these above different methods can differ by an order of mag-
nitude, which could be mainly attributed to the complex het-
erogeneous structure of biochars and the lack of reliable ana-
lytical methods that allow complete characterization of the
biochar continuum in environmental matrices. Traditional
methods for estimating biochar in soils are not considered to
be entirely quantitative and usually lead to an overestimation
or underestimation of biochar content in soil. For example,
measuring benzene polycarboxylic acids (BPCAs) in
biochar-amended soils following nitric acid digestion as
markers for estimating the degree of aromatic condensation
of biochar was adopted in some studies (Glaser et al. 1998;
Wolf et al. 2013; Singh et al. 2014). But, the concentrated
nitric acid digestion consumes a portion of the biochar leading
to an underestimation of soil biochar content by as much as
70 % (Kurth et al. 2006). Even the advanced methods, for
example, using a combination of high-energy ultraviolet
photo-oxidation and NMR spectroscopy to test for a signal
in the aromatic region, were also criticized as both time-
consuming and expensive (Kurth et al. 2006). Thus, method
limitations hamper our interpretation of the ecological func-
tions of biochar, and to find an inexpensive, rapid, and simple
method to quantify biochar in soils and sediment is needed
and would be of great value.
3 The role of biochar in C sequestration
The potential importance of biochar in the global carbon
cycle has intrigued increasing interest recently. Generally,
wildfire produces the largest source of biochar globally; for
example, biochar is present in most boreal forest soils as a
result of naturally recurring wildfires, so that in boreal forest
systems, biochar accounts for 1 Pg of carbon globally (Hart
and Luckai 2013). Biochar produced from fire returning it to
soil can increase the soil organic C pool. For example, a
study carried in Siberian Scots pine forests showed that fire
increased soil C stocks by 16 g kg
−1
or by 40 %. It was
estimated that fire produces biochar as a net terrestrial sink
equivalent to 0.2–0.6 Gt year
−1
atmospheric CO
2
–C; thus,
biochar should be considered as a global C sink of sufficient
importance to be included in the IPPC global C budget
(Santin et al. 2015a,b). However, Billings and Schlesinger
(2015) argued that it should not be included, as it merely
Fig. 2 Interactive layer between biochar, plant root and root-associated
organic matter in profile (a,b) sampled in the largest forest in China
(Daxinanling)
2008 J Soils Sediments (2016) 16:2005–2020
represents the transformation of one form of fixed C (plant
biomass) into another (biochar). In fact, fire converted plant
materials into alkyl and aromatic structures with heterocyclic
macromolecules and small clusters of condensed carbon
with greater C sequestration potential than raw materials
(Czimczik et al. 2003). This recalcitrance ability of biochar
and its quick production process make it much more rapid
than the formation of soil organic matter via the humifica-
tion, which takes decades to become significant (Alexis
et al. 2006). Thus, biochar produced from fire may be an
effective approach to increase C sequestration in a relatively
short time (Lehmann 2007).
3.1 Effects of wildf ire on C turnover
Fire can cause both a substantial loss and gain of C in
forest system. Eckmeier et al. (2007)showedthatgaseous
C losses occurring during fire could be as large as 58.4 t
Cha
−1
. An estimated 79 % of pre-fire vegetation leaf C
was lost as a component of aerosols or gases to the atmo-
sphere (Alexis et al. 2006). Fire-caused C loss also depends
on fire severity and plant species. For example, most of the
pre-burn carbon in parkland vegetation was emitted to the
atmosphere (1300–1500 kg C ha
−1
), and open woodland
ecosystems emitted between 600 and 700 kg C ha
−1
(Imbrozio et al. 2005). Previous studies focused mainly
on the loss of C during and after fire; the estimates of
conversion from plant to biochar and its annual global bio-
char production during fire have been less studied.
Generally, it was reported that conversion rates of plant
material to biochar are between 0.6 and 9 % (Table 1).
Ohlson et al. (2009) estimated about 0.6–2.7 % of plant
material was finally converted into biochar in different
fire-prone ecosystems. Similarly, 0.7–2 % of biomass was
converted into biochar in boreal forests (Hart and Luckai
2013). A much larger range of conversion rate between 1.6
and 8.0 %, with an average of 67.7 kg C ha
−1
, was report-
ed by Barbosa and Fearnside (2005). Another experiment
indicated that 235 kg ha
−1
biochar remained after fire
(Ohlson and Tryterud 2000). Alexis et al. (2006)alsore-
ported that biochar production ranged between 4 and 6 %
of the fire-affected plant biomass. In another study, the
conversion rate of biomass carbon to biochar was consis-
tently 4.8 % (Eckmeier et al. 2007). Variations in biochar
conversion rates are influenced by fire temperature, burning
time, and the type and mass of combustible materials (Hart
and Luckai 2013). Materials with complex carbon struc-
tures usually have higher conversion rates to biochar. In
general, O-alkyl compounds are preferentially lost during
fire, compared to aryl and alkyl compounds, as revealed
by NMR analyses, which indicated a preferential loss of
cellulose components and a relative preservation of lignin
(Alexis et al. 2010). This explains that conifer wood with
high lignin content produced more biochar than grasses
(Demirbas 2001). However, in another study, regardless of
the type of vegetation burned, fires converted about 3 % of
initial ecosystem C to biochar (Forbes et al. 2006), indicat-
ing that differences in plant materials are not the principle
determinant of its conversion efficiency. This is consistent
with biochar properties and its stability potentials as
discussed above.
In addition to plant materials, labile soil organic C is also
frequently converted to more degradation-resistant aromatic
C in boreal forests (Santin et al. 2015a). Santin et al. (2015b)
found that nearly 10 % of mass in forest floor was converted
to biochar, while the corresponding figure with down wood
was 85 %, in a jack pine forest (Pinus banksiana). Unlike
industrially produced biochar, forest-derived biochars are
more complex due to reactions between plant materials
and SOC and nutrients during formation from biomass com-
bustion in the natural soil environment. While most of nu-
trients and organic C will be volatiled and removed from
soil, a small part will be incorporated into biochar during
combustion. How soil nutrients and minerals are involved in
the biochar conversion process during fire awaits systematic
research in field studies.
3.2 Biochar decomposition
Biochar is often considered to be highly stable in soil and
sediments. Large biochar particles which originated from for-
est wildfires may remain in soil for thousands of years
(Czimczik and Masiello 2007). For example, Santos et al.
(2012)applied
13
C-enriched biochar to conduct an incubation
study with sampled temperate forest soils and estimated that
the mean residence time (MRT) ofbiochar was on a centennial
timescale (up to 600 years). However, meta-analysis-based
estimates of turnover time were 88 years on average, which
challenge the concept that biochar can persist in soils for thou-
sands of years (Singh et al. 2012). The processes regulating
biochar decay are poorly understood (Singh et al. 2014).
Biochar decomposition is largely determined by how it was
produced under different environmental conditions. Wildfire
with different combustion conditions (temperatures and oxy-
gen) (Kasin and Ohlson 2013) on different materials in forest
systems provides biochars with unique characteristics which
determine their recalcitrant abilities. Biochar stability was
considered to be dependent on initial biomass type, pyrolysis
temperature, and whether from laboratory incubations or field
studies. For example, leaf-derived biochar produced during
natural vegetation fires does not contribute much to the highly
stable fraction of pyrogenic organic matter compared with
other, more lignified, materials (Alexis et al. 2010). Even with
biochar derived from the same materials, it has both refractory
and dynamic properties, which gives different turnover times.
For example, using a two-pool model approach gave the best
J Soils Sediments (2016) 16:2005–2020 2009
Tab le 1 C balance in post-fire soil within different forest ecosystems
Location Vegetation Type of fire Time since fire Biochar conversion
a
Biochar stock (below
and above ground)
Average decomposition rate
and mean residence time
Reference
Andalusia, Mediterranean Sclerophyllous
forest
Wild fire Annual 1.747 Mg ha
−1
0.029 % yearly Gonzalez-Perez et al.
(2004)
Brazilian Amazonia Shrubs Wild fire –1.8 Mg ha-1 burnt hay.
15.27 to 16.96 %
1.33 to 1.56 Mg ha
−1
na Barbosa and
Fearnside (2005)
Southern region of
North America
Boreal forest Wild fire 14 to 208 years na na 0.23 to 0.28 % yearly Hart et al. (2013)
Northern Sweden Boreal forest Wild fire 1 to 350 years na 0.984 to 2.074 Mg ha
−1
,
average 1.46 Mg ha
−1
na Zackrisson Olle et al.
(1996)
Florida Scrub oak Prescribed fire 10 years 1.75 Mg ha
−1
burnt 4 to 6 % of vegetation
and litter, amounting
to 1.40 Mg ha
−1
na McHenry (2010)
Brazilian Amazonia Seasonal forest Wild fire –0.25 to 4.04 % of living
tree biomass
3.45 ± 2.17 Mg ha
−1
na Turcios et al. (2016)
Canada Boreal forest Wild fire –0.7 to 2.0 % of burnt
biomass in boreal
forest
0.28 to 7.55 Mg ha
−1
Biochar MRT is a few
hundred years
Hart et al. (2013)
Northwest of Zurich Beech-dominated
temperate forest
Lab-produced char
and field trial
10 months 62.45 % of pine wood
feedstock
11.2 ± 2.9 % remaining
after 10 months
applied to soil.
Equivalent to 8.59 to 9.17 % per
month (char loss including
migration, leaching, and
decomposition)
Singh et al. (2014)
Northwest
Territories, Canada
Jack pine Experimental fire na 27.6 % detected by
experimental fire
na na Santin et al. (2015a,b)
Switzerland Pine wood Lab-produced char
and incubation
48 days 49 % of pine wood
feedstock
99.34 to 99.54 % Ranged from 0.46 to 0.66 % in the
first month, no further detectable
decline until incubation ended
Hilscher et al. (2009)
Norway Scots pine Lab-produced char
and field trial
20 months 76.2 to 89.9 % of plant
tissue
94.02 to 98.10 %,
after 20 months
field trial
Ranged from 1.1 to 1.5 % per
month in the first 2 months,
followed by 0.70 to 4.14 %
yearly
Kasin and Ohlson
(2013)
Tuscany, central Italy Pine needles and
wood
Lab-produced char
and incubation
1 month 65.64 % for average
of wood biomass
99.43 to 99.53 % Range 0.47 to 0.57 % in the
first month
Nocentinietal.
(2010)
na not available
a
Since data was incomplete in some papers such as burnt area, pre-fire biomass, etc. to present comparable data, biochar conversion was unified to two types of data, percentage of vegetation biomass and
megagrams per hectare
2010 J Soils Sediments (2016) 16:2005–2020
estimate of biochar decomposition rates, which were 3 years
for a fast-cycling pool within the fresh biochar and 870 years
for a slow-cycling pool in the aged biochar (Singh et al. 2012).
This is consistent with other studies (Luo et al. 2011; Smith
et al. 2010;Kuzyakovetal.2014) with one exception that the
author believes that aged biochar can release higher soluble
and colloidal fractions compared to fresh biochar (Abiven
et al. 2011).
In addition to the refractory and dynamic properties of bio-
char, the mechanisms involved in its protection and loss are
important. After its production, the soil type, mineralology,
climate, and biota all appear to play roles in controlling bio-
char accumulation in the soil C pool (Czimczik and Masiello
2007; Santos et al. 2012; Luo et al. 2013) and the combina-
tions of soil physical, chemical, and biological processes in-
teract in the degradation of biochar, e.g., pH (Luo et al. 2011),
organic matter content (Kuzyakov et al. 2009), and CEC
(Cheng et al. 2006). Soil minerals provide physical and chem-
ical protection to SOM through mineral-associated interac-
tions (Mikutta et al. 2006;Mossetal.2010). If biochar can
interact with soil mineral fractions, how soil mineral compo-
sition influences mineralization rates of biochar raises a sig-
nificant question (Santos et al. 2012). However, very little is
known about this.
Biochar mineralization is stimulated by the presence of
readily mineralizable substrates (Hamer and Marschner
2005). Nocentini et al. (2010) observed a faster rate of biochar
mineralization following addition of glucose. In their experi-
ment, addition of
13
C-labeled glucose increased biochar turn-
over rates, whereas cellulose addition did not. However, the
addition of sugarcane was recently reported to have no signif-
icant effect on overall mineralization of biochar during 7 years
of incubation (Dharmakeerthi et al. 2015). This discrepancy
might depend on whether substrate addition could access mi-
croorganisms which caused co-metabolism, which usually
leads to increases in both SOM and biochar mineralization
(Schmidt et al. 2011). Substrate addition probably increases
the decomposition of the labile component of biochar in the
early stages (Luo et al. 2011), with much lower decomposition
rates in the long term (Kuzyakov et al. 2009). However, re-
search using forest soil or in situ studies are still lacking. This
work is necessary because many substrates, e.g., leaves, are
ubiquitous in forest ecosystems, which links plant species,
SOM, and biochar.
Biochar decomposition is a complex process; it does not
follow a simple model reflecting fire intensity or frequency of
biochar inputs to soil. Some studies report that biochar accu-
mulates in soils, while others report lower-than-expected con-
tributions to stable soil C pools (Czimczik and Masiello
2007). There is still much uncertainty concerning the accumu-
lation of biochar in post-fire forest soils. Recent research on
the fate of biochar in the environment has mainly focused on
its degradation pathways; its accumulation and
transformations in surface soils have been largely ignored
(Santin et al. 2015a,b). Biochar accumulation is determined
not only by its recalcitrant ability or environmental conditions,
but also by its C distribution in different pools, including
particulate C, dissolved organic C leaching, assimilation into
the microbial biomass C pool, etc. All need careful consider-
ation in the context of a biochar–soil system. Only when the
fate of biochar can be accurately traced can biochar accumu-
lation in the complicated soil–biochar system be fully
evaluated.
3.3 Difficulty in estimating biochar mineralization rates
A study collating data from 53 natural biochars found that
differences in biochar mineralization rates could not be
discerned by biochar age, which indicated the difficulty
in quantifying the effects of vegetation fires on global C
cycling (McBeath et al. 2013), reflecting the problems in
estimating biochar decomposition rates in the field.
Reasons for this might be attributed to the following: (1)
Biochar is not a homogenous material with uniform sta-
bility. (2) Estimates of rates of biochar decomposition can
be compromised by the presence of small amounts of
more readily mineralized non-char material that is at-
tached or adsorbed to the char. (3) Biochar decomposes
very slowly, so biochar mineralization in situ cannot be
reliably measured unless isotopic methods are employed.
In addition, most studies cannot distinguish biochar
decomposition and SOC mineralization in a mixture of
biochar and SOC without using isotope techniques. For
example, Wardle et al. (2008b) apparently found an in-
creased SOC mineralization with biochar addition.
However, this finding has been questioned by others as
no account was taken of the discrimination between dif-
ferent pools by using isotopic techniques (Lehmann and
Sohi 2008;Wardleetal.2008a). These factors complicate
estimates of biochar decomposition time (Kane et al.
2010), designed to better understand the link between fire,
biochar, and soil in terms of their potential for carbon
sequestration, both in incubation experiments under con-
trolled laboratory conditions (Liang et al. 2010; Luo et al.
2011) without or with substrates (Nocentini et al. 2010).
An initiation of long-term field experiments (Wardle et al.
2008b; Kasin and Ohlson 2013) is recommended in dif-
ferent forest systems under different climates, by adoption
of techniques such as radiocarbon (
14
C) and stable isotope
(
13
C) analysis and
13
C NMR to address the decomposition
dynamics of biochar.
3.4 Priming effects induced by biochar
Biochar can interact with other forms of organic matter, in-
cluding SOM, for example. This was also observed by Wardle
J Soils Sediments (2016) 16:2005–2020 2011
et al. (2008b) who measured significantly greater SOM min-
eralization following biochar incorporation in a forest soil.
However, this is very controversial (Lehmann and Sohi
2008; Wardle et al. 2008a) and was followed with research
aiming to understand if positive or negative priming effects
occurred with biochar incorporation (Luo et al. 2011;
Zimmermann et al. 2012; Maestrini et al. 2014). The appar-
ent contradiction between the high stability of biochar and
its impact on soil organic matter mineralization rates needs
to be resolved (Woolf et al. 2010). However, most research
is restricted to laboratory studies using soils sampled from
agriculture system. There are only a few studies investigat-
ing forest soil priming effects with biochar (Santos et al.
2012; Bruckman et al. 2015). Santos et al. (2012)observed
initial primed CO
2
in situ in a forest site but no difference
between soil with and without biochar during 15 months.
Another 10-month incubation study with six contrasting bo-
real forest soil types and nine charcoal types (each from
different woody plant species) showed that incorporated bio-
char caused an enhanced loss of total soil organic matter in a
boreal forest soil, with consistent effects among contrasting
charcoal and soil types. They also suggested that biochar-
induced priming effects might affect forest processes that
contribute to the ecosystem C balance and ecosystem func-
tioning, which is important and should not be overlooked
(Pluchon et al. 2016).
4 Biochar effects on soil N dynamics
In the post-fire forest ecosystem, biochar was reported to
have the potential to enhance soil nutrient availability
(Zackrisson et al. 1996; Pietikäinen et al. 2000), and it
plays an important role in regulating N transformations
(Berglund et al. 2004). There has been much work inves-
tigating N cycling in post-fire forest soils containing bio-
char, including nitrification (DeLuca et al. 2006), nitrous
oxide emissions reduction (Clough and Condron 2010).
Since these N transformations may vary with soil type,
biochar may have significant effects on N cycling in dif-
ferent soil systems. Recent studies quantified soil N trans-
formations in forest soils from temperate regions (DeLuca
et al. 2006; Gómez-Rey and Gonzalez-Prieto 2013;
Maestrini et al. 2014) to boreal region (DeLuca et al.
2002) although most recent studies are limited to arable
soils (Joseph et al. 2010; Zhu et al. 2014).
Studies from different post-fire forest systems are given
in Table 2. Soil ammonium concentrations generally in-
creased two to seven times over pre-fire levels following a
fire, and most effects were short lived, being less than
2 years (Chorover et al. 1994). In addition, most studies
showed that soil nitrate pools increased two to five times
after fire, with one exception where soil nitrate increased
44 times after a Mediterranean forest fire (Rapp 1990).
There are conflicting results about whether N mineraliza-
tion rates are enhanced or diminished after fire. Within the
same ecosystem (P. ponderosa), some found increased N
mineralization rates (Kaye and Hart 1998; Deluca and
Zouhar 2000; Choromanska and DeLuca 2002), while
others reported the opposite response (Monleon et al.
1997; Wright and Hart 1997; Deluca and Zouhar 2000;
Grady and Hart 2006; Leduc and Rothstein 2007). It is
concluded that variable responses of total soil N to fire
range from 65 to 275 % compared to pre-fire controls
(Table 2), largely depending on the balance between N
inputs to the soil from ash and canopy litter versus com-
bustion and leaching losses.
Previous work indicates that biochar has the potential
to enhance nitrification, giving high concentrations of
inorganic N, which satisfies plant N demands. A labora-
tory incubation experiment over 158 days measured SOC
mineralization, gross N mineralization, and NH
4
+
and
NO
3
−
contents following the incorporation of ryegrass-
derived biochar in a temperate forest soil. The ammoni-
umwasrapidlytransformedintonitrateandstoredinthis
form throughout the incubation period (Maestrini et al.
2014). Deluca et al.(2006) measured both net and gross
nitrification and concluded that biochar might enhance
the gross rate of nitrification rather than simply decrease
immobilization rates.
Whether this effect of biochar on N mineralization rates
is long term or short term is not clear. Its effect on nitrifi-
cation rates was observed immediately after fire and can be
maintained for many years afterward. Again, the change of
nitrification rate after fire might be from the direct fire or
from the biochar produced. As fire itself could also affect
nitrification rate via changing soil physio-chemical proper-
ties in a relatively long term, it is difficult to distinguish the
fire-induced effects and biochar effects. DeLuca et al.
(2006) found that in Ponderosa pine forests of the Inland
Northwest, the biochar produced by a 12-year-old wildfire
caused greater nitrification rates compared with sites with-
out fire for 75 years. In another forest at Acadia National
Park, which was burnt in 1947, the post-fire biochar with
changed tree species has still maintained higher dissolved
inorganic N (Nelson et al. 2007). In contrast, it was also
reported that fire-induced N responses were only main-
tained for a relatively short time; thereafter, both nitrate
concentration and potential nitrification decreased after
the fire was extinguished (MacKenzie et al. 2004).
The presence of biochar definitely affects the N minerali-
zation pattern in forest systems, and these changes can be
explained by the following mechanisms: (1) biochar may ad-
sorb organic compounds that have negative effects on nitrifi-
cation (Ball et al. 2010). Biochar was shown to be extremely
effective at sorbing phenols, removing over 80 % of phenolic
2012 J Soils Sediments (2016) 16:2005–2020
compounds from solution (MacKenzie and DeLuca 2006).
Also, phenol concentrations were related to biochar combus-
tion temperature. Higher temperatures produced biochar with
lower soluble and total phenol concentrations (Gundale and
DeLuca 2006). Thus, the increased nitrification rate was at-
tributed to the adsorption of soluble phenols, which are known
to inhibit nitrification, on the biochar produced after fire
(Clough and Condron 2010). (2) Fire heats the soil, producing
large amounts of ammonium which is later available for nitri-
fication, although fire appears to reduce stocks of mineraliz-
able N through volatilization and fuel combustion.
Nevertheless, NH
4
+
might not provide an N source for too
long, and N mineralization appears to be limited by rapid
NH
4
+
immobilization or adsorption within the porous struc-
ture of activated biochar (as a surrogate for natural-occurring
fire-produced biochar); thus, nitrification post-fire may be lim-
ited by a lack of available NH
4
+
(MacKenzie et al. 2004). In
addition, the increased CEC (53–538 %) with biochar,
resulting from surface oxidation of aromatic rings with large
numbers of negatively charged sites (Liang et al. 2006), can
enhance N availability while avoiding leaching by adsorbing
NH
4
+
. (3) Nitrification is often limited by an unfavorable en-
vironment, e.g., low soil pH. Ammonia monooxygenase, the
first key enzyme in the nitrification pathway, uses NH
3
as a
substrate rather than NH
4
+
in soils with biochar, since the
existence of NH
3
in soils is pH dependent (available NH
3
decreases dramatically with decreasing pH). Thus, the in-
creased pH caused by biochar can give higher available
NH
3
, providing ammonia monooxygenase with more sub-
strate, and so increase nitrification (Chorover et al. 1994;
Kaye and Hart 1998). We also found a correlation between
pH and nitrification rate (Fig. 3), with an exponential increase
of nitrification rate as pH increased. However, in another
study, without significantly altering bulk soil pH, biochar also
increased nitrification in dry montane forest soils (DeLuca
et al. 2006). (4) Biochar can affect N cycling by stimulating
Table 2 N dynamics after fire in different ecosystems
No. Ecosystem NH
4
+
-N
change
a
(%)
NO
3
−
-N
change (%)
Net mineralization
change (%)
Nitrification
change (%)
Total nitrogen
change (%)
Fire type Reference
1 White fir Increase Increase –Increase 108–152 Prescribed Chorover et al. (1994)
2 Ponderosa
pine–
bunchgrass
196 273 200–300 300–500 116 Prescribed Kaye and Hart (1998)
3Pinus pinea 400 4400 1278 –– Prescribed Rapp (1990)
4 Southwestern
Ponderosa pine
112 150 73 –94 Prescribed Wright and Hart (1997)
5 Pine—pitch pine 156–217 85–385 ––65 Wildfire Groeschl et al.
(1993)
6 Ponderosa pine 109–142 275 370 –275 Prescribed Choromanska and
Deluca (2001)
7 Ponderosa pine 713 192 53–524 –120 Prescribed Deluca and
Zouhar (2000)
8 Ponderosa pine 300 NS Decrease NS –Prescribed Monleon et al. (1997)
9 Jack pine 105 325 Decrease NS NS Wildfire Leduc and
Rothstein (2007)
10 Southwestern
Ponderosa pine
Increase Increase Decrease 130 NS Wildfire/
prescribed
Grady and Hart
(2006)
NS not significant, na not available
a
Value of changes is the ratio (%) between after fire and before fire
Fig. 3 The correlation between pH and nitrification rate. Closed triangle
Lavoie et al. (2010), open triangle Turner et al. (2011), closed circle Ball
et al. (2010), open circle Prommer et al. (2014b), closed square Stark and
Hart (1997), open square Koyama et al. (2010)
J Soils Sediments (2016) 16:2005–2020 2013
microbial soil processes. Ball et al. (2010)demonstratedthat
biochar, produced by wildfire in coniferous-dominated for-
ests, increased soil nitrification rates and nitrifier abundance.
The influence of biochar on N processes also has further
implications in terms of N
2
fixation. Post-fire N
2
fixation may
vary with pre-fire conditions that control post-fire heterogene-
ity. Fire may stimulate asymbiotic N
2
fixation activity by al-
tering environmental factors that modulate rates of fixation,
because fixation activity is positively related to pH and tem-
perature increases, which enhance fixation rates for plants of
legumes. In addition, enhanced biological N
2
fixation could
be attributed to increases in available nutrients, formed during
biochar production, which provides a source of available nu-
trients to plants.
Another property of biochar is to reduce emissions of N
2
O,
which is one of the most important greenhouse gases (GHGs).
However, to our knowledge, there has been no such research
in forest systems, only in pasture systems (Clough and
Condron 2010). A mesocosm experiment was performed,
which found that the combination of biochar (20 t ha
−1
)mixed
with urine (760 kg N ha
−1
) could reduce N
2
O losses, com-
pared to control soil. The mechanisms involved might be due
to the environment changes in soil caused by biochar, such as
soil aeration and pH. Another likely reason might be the en-
hanced adsorption ability of inorganic N biochar, especially
NH
4
+
, thus reducing the inorganic N pool available for subse-
quent nitrification- and denitrification-induced losses of N
2
O
(Clough and Condron 2010).
To our knowledge, few studies have effectively evaluated
the potential for biochar in forest soils to influence N dynam-
ics by specific microbial communities. Genes involved in N
cycling mainly include the following: (i) nifH, responsible for
transforming N
2
to NH
3
,(ii)amoA, responsible for oxidizing
NH
3
,(iii)nirS and nirK, responsible for converting NO
2
to
NO, and (iv) nosZ, responsible for reducing N
2
OtoN
2
.Itis
currently hypothesized that biochar affects N cycling by stim-
ulating certain group of soil microorganisms (Ball et al. 2010).
For example, in a 6-month study, quantitative real-time poly-
merase chain reaction (qPCR) analysis showed that genes in-
volved in nitrogen fixation (nifH) and denitrification (nirS)
had significantly greater abundances in 10 % biochar-
amended treatments compared to the control and were signif-
icantly correlated with soil NO
3
−
and total N (Ducey et al.
2013). Oxidation of ammonia is the first step in nitrification;
both ammonia-oxidizing bacteria (AOB) and ammonia-
oxidizing archaea (AOA) play an important role in the ammo-
nia oxidation process in various environments (Fig. 4). A
greater abundance of AOB was attributed to the greater nitri-
fication in soils, which were exposed to fire 12 years previ-
ously compared with control soils (DeLuca et al. 2006). We
also show that with increasing pH, the ratio of AOA to AOB
decreased (Fig. 4), indicating that AOB mainly dominate acid
soils in terms of nitrification.
5 Changes in microbial abundance and shifts
in microbial community structure
Our current understanding of how biochar influences forest
soil processes is mainly limited by insufficient results on how
biochar affects microbial properties. There has been much
work on the effects of biochars on microorganisms in arable
soils (Steiner et al. 2008;Haleetal.2013), but much less on
how biochar affects microbial abundance and the shifts in
microbial communities in forest systems (Zackrisson et al.
1996). The heating during wildfire would kill most microor-
ganisms on the surface soil layer, thus decreasing soil micro-
bial biomass immediately. However, the short-term fire effects
disperse rapidly after wildfires and the biochar exerts signifi-
cant effects on soil microorganisms afterward, although the
biochar might behave differently in the short and long terms.
For example, biochar might adsorb and stabilize organic com-
pounds following their formation and release due to the fire
and prevent volatilization and leaching losses. This could ben-
efit microbes in the early stages.
Pietikäinen et al. (2000) found a reduction in microbial
biomass and a higher specific growth rate with biochar in
the underlying humus after fire. There are also reports that
biochar had no effects on soil microbial biomass (Santos
et al. 2012). In addition to the total amount, the microbial
community structure was changed in biochar-amended soil.
While the overall microbial diversity decreased upon biochar
amendment, there were increases in specific taxa during incu-
bation of biochar-amended soils. A shift in favor of certain
type of microorganisms was also reported (Hale et al. 2013).
There were increases in the ratio of Gram-positive bacteria/
fungi and lower ratios of Gram-negative/Gram-positive bac-
teria (Noyce et al. 2015). Another study investigated the soil
microbial responses to biochar addition in a temperate forest
Fig. 4 The correlation between pH and the ratio of AOA to AOB. Closed
triangle Prommer et al. (2014), open triangle Song et al. (2014), closed
circle He et al. (2015), open circle Harter et al. (2014), open square Bai
et al. (2015), closed square unpublished data
2014 J Soils Sediments (2016) 16:2005–2020
(Haliburton) soil, and the author found that Gram-positive and
Gram-negative bacteria and actinomycetes were lower than
controls in the first 16 weeks, then increased until 24 weeks,
suggesting a gradual microbial adaptation to altered soil con-
ditions (Ding et al. 2013). It was concluded that bacteria can
rapidly adapt to a biochar-amended soil environment
(Mitchell et al. 2015). The DNA sequencing of taxa further
revealed that biochar mainly increased relative abundance of
Actinobacteria and Gemmatimonadetes (Khodadad et al.
2011). In contrast, biochar addition was reported to have little
effects on soil community structure as measured by PLFAs
(Santos et al. 2012). It was also found that the adaptation of
microbial groups to biochar in forest soil was less pronounced
than in arable soil (Wardle et al. 2008a).
The differences in the microbial biomass C pool and its
structure following biochar addition have been explained by
differences in biochar type and soil type (Pietikäinen et al.
2000;Kolbetal.2009; Kuzyakov et al. 2009; Warnock
et al. 2010), but there are some contradictory results in the
literature. The effects of addition of biochars on soil microor-
ganisms have also been reported to be largely dependent on
soil type (Khodadad et al. 2011). However, in a similar study
in temperate soils, biochar addition affected microbial bio-
mass, microbial activity, and nutrient availability in relatively
similar ways in four distinct soil types (Mollisol, Alfisol,
Entisol, and Spodosol) (Kolb et al. 2009). Biochar types, rath-
er than soil types, seem to be the major factors affecting the
microbial community. Steinbeiss et al. (2009) reported that
glucose-derived biochar hardly changed the composition of
the soil microbial community, while yeast-derived biochar
strongly promoted fungi.
The mechanisms involved between biochar and soil micro-
bial interactions still remain largely unknown, and their rela-
tionship to biochar chemical and physical properties is a pri-
ority research area (Liang et al. 2010; Lehmann et al. 2011).
The most likely mechanism involved is considered to be in-
creased C and nutrient availability in biochar-amended soil
(Warnock et al. 2007;Kolbetal.2009; Thies and Rillig
2009), although biochar itself is believed to provide a recalci-
trant C source for microbes (Pietikäinen et al. 2000). In gen-
eral, biochar is a less suitable substrate for bacteria, as they
mainly use easily metabolizable and labile substances, such as
simple sugars, starch, fats, and proteins as sources of nutrition,
while fungi, regarded as K-strategies, can use more complex
materials (Fontaine et al. 2003). Some microorganisms have
been reported to be adapted to living on charred wood from
forest fires and are potential utilizers of biochar (Pietikäinen
et al. 2000). However, direct evident is still inadequate; only a
few studies have used isotope techniques to trace biochar C
sources intothe biomass C pool, and most work is not in forest
soils (Kuzyakov et al. 2009; Luo et al. 2013). In only one
study carried out in forest systems, it was found that the (main-
ly plant) materials from which biochar was derived
determined which specific group of microorganisms used it
as a C and energy source (Steinbeiss et al. 2009). Wardle et al.
(2008a) stated that microorganisms in forest soils seemed to
be less specialized in utilizing different carbon sources. The
large surface area and porous structure of biochar create a
habitat for microbial colonization (Pietikäinen et al. 2000;
Ezawa et al. 2002; Saito and Marumoto 2002; Warnock
et al. 2007; Thies and Rillig 2009). Although the surface area
may be up to 3.6 × 10
9
m
2
ha
−1
with large numbers of reactive
sites (Zackrisson et al. 1996), direct evidence, e.g., SEM im-
ages, for example, is still rare. In addition, fire produces large
quantities of mineral nutrients in the form of ash and this
might influence the availability of nutrients by changing pH
values, even though this effect might only persist for months.
Forest soils are commonly acidic, and the increase in soil pH
following a fire may change the microenvironment for micro-
organisms and thus play a vital role in controlling soil micro-
bial properties (Lehmann et al. 2011).
6 Perspectives
Biochar has great potential in plant productivity and similar
improvements to soil. Although there is a large publication
database for arable systems and quartz-based artificial soils
(Ding et al. 2013), there is a very limited one for forest soils.
Most published research related to forest biochar was carried
out before 2010 (Zackrisson et al. 1996). This limits our abil-
ity to predict the effects of biochar in the soils of forest sys-
tems. The following general recommendations for future re-
search can be made to solve some remaining questions related
to biochar in forest system.
6.1Biocharhumificationprocess
For biochar to meaningfully increase the total pool of humic
substances, it must integrate fully in terms of chemical, phys-
ical, and biological properties. The extent of this integration, if
any, still needs investigation. Some materials in biochar are
believed to be involved in increasing the proportion of humic
acid-like materials in the humic substances of soil organic
matter or compost. However, further research is required to
determine how biochar is transformed in forest soil with time.
For example, carbon from the degradation of lignite was
found to be part of the humic fraction in lignite-rich mine
soils, which may indicate that lignite is oxidized during bio-
degradation and incorporated to the humic fraction. A similar
approach can be adopted in biochar research. The biochars
from different forest fires in different years could provide data
on their spatial distribution, combined with analysis of phys-
ical and chemical properties and especially the degree of
weathering. This would help us to deepen our understanding
J Soils Sediments (2016) 16:2005–2020 2015
of the role of biotic and abiotic factors involved in biochar
humification processes.
6.2 Functional microorganisms driving nitrogen
transformation
Considerable knowledge gaps still exist in terms of under-
standing the precise mechanisms through which biochar influ-
ences soil N transformations. To better understand how bio-
chars affect N transformation in soils after forest fires, three
key processes of N transformation, i.e., N
2
fixation, nitrifica-
tion, and denitrification, have been measured. Combined with
in situ field sampling and laboratory incubation experiments,
by using RT-PCR, DNA-SIP, and high-throughput sequencing
technology, future work should aim to reveal how biochar
affects soil microbial abundance and genetic diversity of func-
tional genomics related to nitrogen transformation, like, nitro-
gen fixation (nifH gene), oxidation of ammonia (amoA gene),
nitrite-oxidizing (NOB), and denitrifying microorganisms
(narG,nirS/K,norB,andnosZ genes), to clarify biochar in-
fluence on N transformation and its mechanisms involved. We
also need to understand the response mechanisms of the sur-
face processes to global change and establish a theoretical
basis for biochar strategies to mitigate global warming.
6.3 Interactions of biochar with soil microbial
communities and plants
In the presence of biochar, plants usually show better develop-
ment and performance (Atkinson et al. 2010), probably related
to plant growth-promoting rhizobacteria and their interactions
with biochar (Lehmann et al. 2011; Luo et al. 2013; Philippot
et al. 2013). The interface between biochar and soil has unique
properties, distinct from both the internal biochar and the bulk
soil, analogous to the rhizosphere. Luo et al. (2013)termedthis
unique area the Bcharsphere,^where the biochar surface inter-
acts with the surrounding soil, plant roots, and soil organic
matter. In soils with a long fire history, biochar is often distrib-
uted in the deep layer, in between mineral soil and surface
organic soil (Fig. 1a). It is interesting to know how biochar
interacts with both mineral soil and surface organic soil, while
biochar has distinct properties in terms of pH, surface area, and
aeration (Fig. 5). It is recently acknowledged that soil organic
matter mineralization is governed mainly by the accessibility of
microorganisms to C sources (Schmidt et al. 2011; Dungait
et al. 2012). Thus, whether the charsphere enables microorgan-
isms to have easier access to C adsorbed on surface would be
interesting to investigate, since C adsorption and microorgan-
ism colonization have been widely reported (Lehmann et al.
2011; Luo et al. 2013). However, this remains largely un-
known, and new techniques, like, NanoSIMS, are required to
understand the interactions involved. In addition, in the biochar
and the plant–soil system, how biochar could affect soil organic
C and plant root in the plant–soil–biochar system remains a
research frontier (Weng et al. 2015). For example, it is difficult
to apply methods using classic two stable isotopes (
13
Cand
12
C) to partitionate three C sources. A dual-isotope approach
wasusedtopartitionatesoilCO
2
emissions derived from min-
eralization of SOC, added biochar, and root respiration
(Whitman and Lehmann 2015). Very few studies have ad-
dressed the relationship between biochar and plants in the nat-
ural forest ecosystem where various factors confound plant
responses and biochar effects (Makoto et al. 2011). Also, we
have little knowledge on the interactions between biochar and
soil animals such as worms, termites, protozoa, etc., in forest
systems. This requires further investigation.
Fig. 5 Schematic overview of
charsphere in between mineral
and organic layer (Luo et al.
2013)
2016 J Soils Sediments (2016) 16:2005–2020
6.4 Combined long-term field and short-term laboratory
experiments
The effects of biochar, with its aromatic structure, on soil
processes can persist for years following biochar addition.
The fundamental mechanisms by which biochar affects the
function of forest ecosystems during these longer timescales
are still poorly understood. The application of biochar to forest
ecosystems has received little attention (Mitchell et al. 2015),
and most studies on biochar in soil have been limited to short-
term laboratory incubations using relatively small-scale sam-
pling approaches. For example, the mechanisms of biochar
degradation derived from laboratory experiments need to be
tested to understand their relative significance in different for-
est soil and climate using both field and laboratory studies
(Czimczik and Masiello 2007). Thus, it is necessary to estab-
lish long-term field experiments in forest soils receiving bio-
char application to improve our understanding of its long-term
impact in forest ecosystems. Meanwhile, laboratory experi-
ments should be included as it would be easier to adopt some
new techniques targeting certain questions which field exper-
iments hardly answer. For example, fire largely sterilizes the
soil surface and possibly some slightly deeper layers, depend-
ing on its intensity, and greatly increases the availability of
nutrients from soil minerals and soil organic matter; once the
fire is over, biochar interacts with these processes and the use
of isotopic techniques can help separate fire effect from bio-
char effects before and after fire events.
Acknowledgments This study was supported by the National Science
Foundation of China (41520104001, 41301250), the National Basic
Research Program of China (2014CB441003), and the Fundamental
Research Funds for the Central Universities in China.
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