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GCB Bioenergy. 2021;13:1731–1764.
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wileyonlinelibrary.com/journal/gcbb
Received: 30 January 2021
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Revised: 15 June 2021
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Accepted: 18 June 2021
DOI: 10.1111/gcbb.12885
RESEARCH REVIEW
How biochar works, and when it doesn't: A review of mechanisms
controlling soil and plant responses to biochar
StephenJoseph1,2,3,4
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Annette L.Cowie3,5
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LukasVan Zwieten6,7
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NanthiBolan7,8,9
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AliceBudai10
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WolframBuss11
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Maria LuzCayuela12
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Ellen R.Graber13
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James A.Ippolito14
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YakovKuzyakov15,16,17
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YuLuo18
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Yong SikOk19
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Kumuduni N.Palansooriya19
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JessicaShepherd20
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ScottStephens21
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Zhe (Han)Weng22,23
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JohannesLehmann24
1School of Materials Science and Engineering, University of NSW, Kensington, New South Wales, Australia
2Institute of Resources, Ecosystem and Environment of Agriculture, and Center of Biochar and Green Agriculture, Nanjing Agricultural University, Nanjing,
China
3School of Environmental and Rural Science, University of New England, Armidale, New South Wales, Australia
4ISEM and School of Physics, University of Wollongong, Wollongong, New South Wales, Australia
5New South Wales Department of Primary Industries, Armidale, New South Wales, Australia
6New South Wales Department of Primary Industries, Wollongbar, New South Wales, Australia
7Cooperative Research Centre for High Performance Soil (Soil CRC), Callaghan, New South Wales, Australia
8School of Agriculture and Environment, The University of Western Australia, Perth, Western Australia, Australia
9School of Engineering, College of Engineering, Science and Environment, Callaghan, New South Wales, Australia
10Norwegian Institute of Bioeconomy Research, Division of Environmental and Natural Resources, Ås, Norway
11Research School of Biology, Australian National University, Canberra, Australian Capital Territory, Australia
12Department of Soil and Water Conservation and Waste Management, CEBAS- CSIC, Murcia, Spain
13Institute of Soil, Water and Environmental Sciences, The Volcani Center, Agricultural Research Organization, Rishon LeTzion, Israel
14Department of Soil and Crop Sciences, Colorado State University, Fort Collins, Colorado, USA
15Department of Soil Science of Temperate Ecosystems, Dept. of Agricultural Soil Science, University of Göttingen, Göttingen, Germany
16Agro- Technological Institute, RUDN University, Moscow, Russia
17Institute of Environmental Sciences, Kazan Federal University, Kazan, Russia
18Institute of Soil and Water Resources and Environmental Science, Zhejiang Provincial Key Laboratory of Agricultural Resources and Environment,
Zhejiang University, Hangzhou, China
19Korea Biochar Research Center, APRU Sustainable Waste Management Program & Division of Environmental Science and Ecological Engineering, Korea
University, Seoul, South Korea
20University of Edinburgh School of Geosciences, Edinburgh, UK
21New South Wales Department of Primary Industries, Parramatta, New South Wales, Australia
22School of Agriculture and Food Sciences, The University of Queensland, St. Lucia, Queensland, Australia
23Department of Animal, Plant & Soil Sciences, Centre for AgriBioscience, La Trobe University, Melbourne, Victoria, Australia
24Soil and Crop Sciences, School of Integrative Plant Science, Cornell University, Ithaca, New York, USA
This is an open access article under the terms of the Creat ive Commo ns Attri bution License, which permits use, distribution and reproduction in any medium, provided the original
work is properly cited.
© 2021 The Authors. GCB Bioenergy published by John Wiley & Sons Ltd.
Stephen Joseph and Annette L. Cowie are to be considered joint first authors.
1732
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JOSEPH Et al.
Correspondence
Annette L. Cowie, New South Wales
Department of Primary Industries,
Armidale, NSW 2351, Australia.
Email: Annette.cowie@dpi.nsw.gov.au
Funding information
Program of Competitive Growth of Kazan
Federal University and the “RUDN
University program 5- 100”; La Trobe
University, Grant/Award Number: LTU
SFWE RFA 2000004349
Abstract
We synthesized 20years of research to explain the interrelated processes that deter-
mine soil and plant responses to biochar. The properties of biochar and its effects
within agricultural ecosystems largely depend on feedstock and pyrolysis conditions.
We describe three stages of reactions of biochar in soil: dissolution (1– 3weeks); re-
active surface development (1– 6months); and aging (beyond 6months). As biochar
ages, it is incorporated into soil aggregates, protecting the biochar carbon and promot-
ing the stabilization of rhizodeposits and microbial products. Biochar carbon persists
in soil for hundreds to thousands of years. By increasing pH, porosity, and water
availability, biochars can create favorable conditions for root development and micro-
bial functions. Biochars can catalyze biotic and abiotic reactions, particularly in the
rhizosphere, that increase nutrient supply and uptake by plants, reduce phytotoxins,
stimulate plant development, and increase resilience to disease and environmental
stressors. Meta- analyses found that, on average, biochars increase P availability by a
factor of 4.6; decrease plant tissue concentration of heavy metals by 17%– 39%; build
soil organic carbon through negative priming by 3.8% (range −21% to +20%); and
reduce non- CO2 greenhouse gas emissions from soil by 12%– 50%. Meta- analyses
show average crop yield increases of 10%– 42% with biochar addition, with greatest
increases in low- nutrient P- sorbing acidic soils (common in the tropics), and in sandy
soils in drylands due to increase in nutrient retention and water holding capacity.
Studies report a wide range of plant responses to biochars due to the diversity of bio-
chars and contexts in which biochars have been applied. Crop yields increase strongly
if site- specific soil constraints and nutrient and water limitations are mitigated by
appropriate biochar formulations. Biochars can be tailored to address site constraints
through feedstock selection, by modifying pyrolysis conditions, through pre- or post-
production treatments, or co- application with organic or mineral fertilizers. We dem-
onstrate how, when used wisely, biochar mitigates climate change and supports food
security and the circular economy.
KEYWORDS
carbon sequestration, GHG mitigation, heavy metals, priming effect, resilience, rhizosphere
processes, soil carbon
1
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INTRODUCTION
Biochar is produced by thermal transformation of organic
matter in an oxygen- limited environment. Research interest
in biochar has grown markedly since 2000 (Figure S1), stim-
ulated by early studies of Terra Preta soils in the Amazon that
indicated potential for biochar amendment to simultaneously
improve a broad range of soil properties and thus increase
agricultural yields, while also contributing to climate change
mitigation (Glaser et al., 2002; Lehmann et al., 2006).
A wide range of biochar types produced from feedstocks
including woody residues, crop straw, animal manures, sew-
age sludge, and food wastes are pyrolyzed at temperatures
(highest treatment temperature, HTT) ranging from around
350°C to over 750°C. Biochar properties vary widely, de-
termined largely by feedstock, HTT, and residence time at
HTT, as well as treatments applied before and after pyrolysis
(Schimmelpfennig & Glaser, 2012). A review of 5400 stud-
ies (Ippolito et al., 2020) found that wood- based feedstocks
generally produced biochars with the highest surface area,
straw- based feedstocks gave the highest cation exchange ca-
pacity (CEC), and manure feedstocks produced biochars with
the highest N and P content. HTTs above 500°C produced
biochars that were more persistent in soil, with higher ash
contents and pHs.
Biochar trials have used a wide range of application rates
and formulations (Text S1; Table S1; Figure S2; Figure S3).
Higher rates (10– 50Mgha−1) have commonly been applied
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JOSEPH Et al.
where low- nutrient biochar is used as a soil conditioner to im-
prove bulk soil chemical and physical properties, while lower
rates (<1Mgha−1) have been used as a nutrient carrier to
increase fertilizer use efficiency and decrease nutrient losses,
and in mechanized planting (Table S1). Economic analyses
suggest that formulations combining biochar with fertilizer
(biochar compound fertilizer [BCF]), applied at low rates, are
likely to be the most cost- effective approach for broadacre
cropping in higher income countries (Robb et al., 2020).
Studies report a wide range of effects of biochars on phys-
ical, biological, and chemical soil properties and functions,
and on plant growth. Reviews and meta- analyses show that
biochar generally lowers soil acidity and increases buffer-
ing capacity; increases dissolved and total organic C, CEC,
available nutrients, water retention, and aggregate stability;
and reduces bulk density (El- Naggar et al., 2019; Lehmann &
Joseph, 2015). Biochar can increase microbial activity, accel-
erate nutrient cycling, and reduce leaching and volatilization
of nitrogen (Lehmann & Joseph, 2015).
In terms of plant performance, biochars can affect seed
germination, plant growth, flowering, resistance to disease,
and acclimation to abiotic stresses. Many studies report that
biochar increases plant productivity, with an average yield
increase of 10%– 42% (Table 1), although negative effects
have also been recorded (Jeffery et al., 2017; Macdonald
et al., 2014; Ye et al., 2020). Studies reporting positive re-
sponses have commonly used biochar application rates of
5– 20Mgha−1 (Table 1); however, applications of biochar–
fertilizer mixes at low rates (<1Mgha−1 biochar) have also
increased yields, particularly when applied as a band near the
seed (Table S1). The effects of biochar on crop yields are
discussed further in Section 4.
Besides agronomic benefits, biochar contributes to cli-
mate change mitigation: Biochar C persists in soil for one
to two orders of magnitude longer than unpyrolyzed organic
residues, providing long- term C sequestration when applied
to soil. In addition, biochar can increase soil C levels by de-
creasing mineralization of existing soil organic matter (SOM;
Wang et al., 2016) and newly added plant C (Weng et al.,
2017). Furthermore, biochar can reduce emissions of the
greenhouse gases (GHGs), nitrous oxide and methane (Van
Zwieten, Kammann, et al., 2015).
The large body of literature that has accumulated over
the last two decades has greatly increased our observational
database of the effects biochar can have on soil properties
and crop performance. In- depth mechanistic studies have
brought focus to the importance of the rhizosphere in these
effects. The objectives of this review are to synthesize the last
20years of research on biochar to elucidate the underlying
biochar– soil– plant processes, and mechanisms that lead to
plant responses to biochar, and to provide recommendations
for optimizing the use of biochar to increase plant yield, soil
health, and climate change mitigation.
We first describe biochar– soil– plant interaction mecha-
nisms, focusing on rhizosphere processes and implications
for plant growth, concentrating on biochar applied to annual
crops. Use of biochar in annual crops has been the most com-
monly studied application to date and is anticipated to be the
most widespread future application of biochar. Subsequent
sections review the implications of biochar for food secu-
rity, climate change mitigation, and the role of biochar in the
circular economy. We conclude with a summary of key pro-
cesses, knowledge gaps, and recommendations for optimal
biochar use.
2
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MECHANISMS OF BIOCHAR
EFFECTS ON SOIL AND PLANTS
We consider the interactions between biochar, soil and plants
in the context of the annual crop cycle:
- Stage 1: Short- term (1– 3 weeks) reactions of biochar
in soil, and effects on seed germination and seedlings
- Stage 2: Medium- term (1– 6months) creation of reactive
surfaces on biochar, effects on plant growth and yield
from seedling to harvest
- Stage 3: Long- term (>6months) interactions as biochar
“ages” in soil, and its effect on subsequent crop cycles.
Biochar is commonly applied at sowing or 1– 3weeks be-
fore sowing. Mechanisms involved when biochar is applied
in conjunction with mineral and/or organic fertilizers, and as
a BCF comprising biochar, fertilizer, minerals (e.g., gypsum,
dolomite, diatomite, rock phosphate) and binder (e.g., clay,
starch) are examined.
2.1
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Stage 1: Short- term reactions
(1– 3weeks)
2.1.1
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Biochar reactions in soil
Chemical effects
The general properties of biochars are described in Text S2.
After application to soil, water entering biochar pores dis-
solves soluble organic and mineral compounds on biochar
outer and inner surfaces (Figure 1). These solutes increase
dissolved organic carbon (DOC), cations, and anions in the
soil solution (Silber et al., 2010), which increases the elec-
trical conductivity and pH and reduces Eh (Joseph et al.,
2015) The extent of changes in soil solution composition
depends on the specific biochar and soil (Mukherjee &
Zimmerman, 2013; Schreiter et al., 2020). Release of DOC
and nutrient ions from biochar (Kim et al., 2013) is rapid
over the first week and much slower over the following
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JOSEPH Et al.
TABLE 1 Summary of meta- analyses of yield response to biochar, and synthesis of findings
Study
Number
of studies
included Notes
Biochar dose
giving optimal
response (t/ha)
Crop factor
reported
Grand mean
change
% Synthesis of key findings
Jeffery et al. (2011) 23 First meta- analysis of biochar
effects on yield, using pot
and field studies.
100 Crop yield 10 • Greatest benefit in acidic and pH neutral soils
suggesting liming effect as key driver
• Greater effect in coarse- textured soils suggesting
improved water and nutrient availability
• Poultry litter feedstock showed the greatest positive
benefit
Biederman and Harpole
(2013)
114 Yield response presented
for fertilized biochar
treatment vs. fertilized
control
ns Crop yield 42 • Improved plant tissue P and K concentrations
compared to fertilizer alone
• Grass and manure feedstocks most effective
especially at higher temperatures due to increased
liming effect
• Application rate was not a good predictor due to
variable responses arising from different interactions
Crane- Droesch et al. (2013) 84 Predicted yield response
based on the application
of 3Mg ha−1 biochar
3 Crop yield 10 • Soil CEC and SOC content are predictors of yield
response
• Greatest benefits in lowest- potential agricultural
areas, which are predominantly found in the humid
tropics
• Benefits to yield increased yearly up until year 4 after
application
Liu et al. (2013) 103 <10– 20 Crop yield 11 • Greatest benefit in acidic soils (pH<5)
• Greatest benefit in sandy soils, followed by clay, then
silt or loam- textured soils
• Manure feedstock generally most effective, followed
by wood, then crop residue
Thomas and Gale (2015) 17 Study focused on responses
of trees
NA Tree biomass 41 • Greater positive effect in tropical than in temperate
systems, and in angiosperms than conifers
• Limited number of studies exist but authors suggest
significant opportunities during reforestation
Jeffery et al. (2017) 111 NA Crop yield 13 • An average 25% yield increase in the tropics with
liming effect and fertilizer value as the key driver
• Biochars containing higher nutrient contents had
greater benefit to yield than lower nutrient biochars
• The authors stressed the need to understand the
constraint that is being addressed by biochar
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JOSEPH Et al.
Study
Number
of studies
included Notes
Biochar dose
giving optimal
response (t/ha)
Crop factor
reported
Grand mean
change
% Synthesis of key findings
Xiang et al. (2017) 136 Analysis focused on below
ground effects
NA Root biomass 32 • No change in root N concentration but significantly
increased root P concentration
• Benefits were greater for legumes and resulted in
increased nodulation
• Higher temperature biochars had a greater effect and
HTT was a more important indicator than feedstock
type
Awad et al. (2018) 50 Study focuses on rice
production
1– 10 Crop yield 16 • Greatest benefits observed in very acidic soil
• No significant differences in effects with biochar rate
• Soil texture did not have a major role in determining
yield effect
Dai et al. (2020) 153 NA Crop yield 16 • Liming, improved soil physical properties,
and increased nutrient use efficiency were key
mechanisms resulting in positive effects
• Interactions between soil properties and biochar
properties were key to delivering positive effects
• Biochars with high ash content (e.g., higher nutrient
content) applied into sandy and/or acidic soils likely
to give greatest benefits.
Ye et al. (2020) 56 Comparison with unfertilized
control
<5 Crop yield 30 • The study separately assessed biochar response in
comparison with fertilized and unfertilized controls,
and showed that benefits to yield are additive to the
fertilizer effect
• Soils with CEC<100mmolckg−1 showed the
greatest positive response
• Soils with SOC≤20gkg−1 showed the greatest
positive response
• Crops grown on soils with pH≤6.5 always had a
positive yield response to biochar addition
Comparison with
fertilizedcontrol
5– 10 Crop yield 10
Abbreviations: CEC, cation exchange capacity; NA, not applicable/ not investigated; ns, no response detected; SOC, soil organic carbon.
TABLE 1 (Continued)
1736
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JOSEPH Et al.
weeks (Mukherjee & Zimmerman, 2013). Initial rapid dis-
solution can occur via dissolution of salts, ion exchange,
submicrometer particle detachment, and preferential dis-
solution at crystal imperfections (Wang et al., 2020). After
the initial rapid dissolution stage, continued dissolution is
faster in acidic (Silber et al., 2010) and low- nutrient soils
(Wang et al., 2020).
When biochar is applied in the form of BCF that com-
bines biochar, minerals, and N and P compounds (e.g., urea,
ammonium sulfate, diammonium phosphate), the physical
and chemical reactions that occur during the production of
the granules slow the rate and extent of dissolution of N
compounds compared with dissolution of mineral fertilizers
(Chen et al., 2017; Shi et al., 2020).
Fresh biochar typically has a low CEC, as the high tem-
peratures during pyrolysis reduce the concentration of func-
tional groups (e.g., – OH, – COOH, – CH, and – C=O). CEC of
biochar is more difficult to measure than CEC of soils, due to
its pH- dependent variable charge properties and the presence
of soluble salts (Graber et al., 2017; Munera- Echeverri et al.,
2018). Using methods considered suitable for biochar, CEC
ranges from approximately 50 to 200mmol kg−1 (Graber
et al., 2017; Mitchell et al., 2013), and anion exchange ca-
pacity (AEC) is typically also less than 200 mmol kg−1
(Lawrinenko et al., 2017). As CEC of fresh biochar is rel-
atively low compared with CEC of many soil components,
applying biochar typically does not increase the soil CEC
immediately (Kharel et al., 2019). However, the CEC and
FIGURE 1 Summary of the processes that occur when biochar is applied to soil, based on two modes of application: (left) biochar and
fertilizer applied together and incorporated through the soil prior to sowing, and (right) biochar compound fertilizer (BCF) comprising biochar
mixed with fertilizer, minerals and a binder, granulated, applied to the soil as a band near the seed. (a) Stage 1: dissolution of biochar, interactions
with seedlings; (b) Stage 2: reactive surface development on biochar, interactions with growing plants. RC, resistor and capacitor in parallel
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JOSEPH Et al.
AEC of biochar increase over time as additional functional
groups form on biochar surfaces (see Section 2.3), increasing
its ability to sorb and retain cations and anions (Hagemann,
Joseph, et al., 2017; Hagemann et al., 2017; Rechberger et al.,
2017; de la Rosa et al., 2018; Wang et al., 2019).
Low- temperature biochars (HTT<450°C) and biochars
produced in facilities with incomplete separation of pyrolysis
vapors (Buss & Mašek, 2014; Buss et al., 2015) generally have
higher contents of water- soluble organic compounds, particu-
larly low molecular weight neutrals (alcohols, aldehydes, ke-
tones, phenolics, karrikins), polyphenols/polyphenolic acids,
and complex macromolecules, whereas high- temperature
biochars (HTT >450°C) have relatively lower levels of water-
soluble compounds that are dominated by low- molecular
weight acids and low- molecular weight neutrals (Graber
et al., 2015; Reynolds et al., 2018; Taherymoosavi et al.,
2018). Low- temperature biochars can be hydrophobic ini-
tially due to accumulation of aliphatic compounds in pores
and on the surface; such compounds are usually lost during
pyrolysis at higher temperatures. Hydrophobicity can inhibit
water uptake by biochar particles (Gray et al., 2014), but this
effect dissipates over time.
Most biochars are alkaline, with acid- neutralizing capac-
ity up to 33% of agricultural lime (Van Zwieten, Kimber,
Morris, Chan, et al., 2010) due to their carbonate, oxide, and
hydroxide content. Biochar is a reductant, and therefore low-
ers soil redox potential (Joseph et al., 2015). An exception
is flooded rice soils, where biochar application can increase
Eh due to the release of O2 from roots. Chew et al. (2020),
Joseph et al. (2015), and Pignatello et al. (2017) detail the
range of reactions that can take place on the external surfaces
and in the pores of biochar (see also Section 2.2). Except in
flooded soils, oxygen will diffuse into the pores and react
with redox- active organic molecules (e.g., quinones; Yu &
Kuzyakov, 2021) and minerals, particularly Fe and Mn. In
acid soils, excess H+ reacts with basic minerals such as cal-
cite and dolomite present within the C lattice of the biochar
(Amonette & Joseph, 2009).
Biochars (especially those made at >400°C) can have a
high content of free radicals, which can lead to the forma-
tion of reactive oxygen species (Pignatello et al., 2017; Ruan
et al., 2019; Yu & Kuzyakov, 2021) and strongly accelerate
oxidation reactions. This acceleration leads to oxidation not
only of biochar itself but also of SOM and plant residues (Du
et al., 2020) and is especially intensive in soils with fluctuat-
ing water level (Merino et al., 2020) or with high content of
iron (oxyhydr)oxides (Merino et al., 2020; Yu & Kuzyakov,
2021).
Physical effects
Biochars commonly increase soil water holding capacity,
particularly in coarse- textured soils, decrease bulk density,
and increase porosity, with greater effects observed at rates
exceeding 40 Mg ha−1 (see Section 3; Quin et al., 2014).
Biochar can also impact water infiltration into soils, for ex-
ample, moderating the reduction in infiltration rate that oc-
curs during high- intensity rainstorms in soils prone to surface
sealing, as seen at 2% w/w by Abrol et al. (2016). Reduced
sealing leads to lower runoff and erosion rates. The effects
were attributed to a biochar- related increase in soil solution
Ca and decrease in Na, leading to decreased sodium adsorp-
tion ratio (Abrol et al., 2016).
Biochar particles have low density and are easily crushed
(Abdullah & Wu, 2009). Cultivation and ingestion by soil
fauna result in fragmentation and fracturing, creating very
small particles (approximately <100µm). These small parti-
cles are more mobile and can have higher reactivity, surface
charge, radical content (Das et al., 2020; Yu & Kuzyakov,
2021) and surface area than larger particles (Yang et al., 2020),
which can increase reactivity and nutrient availability (Wang
et al., 2020). High mineral ash biochars and engineered bio-
chars used in BCF generally contain high quantities of small
mineral particles <100µm (especially silica, alumina, Fe/O
and CaCO3, CaHPO4, and Mg compounds) in or on the C
matrix that are easily fragmented from the biochar and are
mobile in soil.
2.1.2
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Effects on seed germination and early
seedling growth
Reported impacts of biochar on germination and seedling
growth range from inhibition to stimulation. Hormesis is
commonly observed, that is, high rates of biochar can have a
detrimental effect, while low rates can be stimulatory. Below
we discuss the mechanisms likely to contribute to the range
of effects on seed germination and early seedling develop-
ment reported in the literature.
Seed germination begins with water imbibition and ends
when the radicle emerges from the seed coat. The following are
the main factors that determine whether biochar impacts seed
germination: (i) release of salts from biochar to the soil solution;
(ii) release of phytotoxins; (iii) release of germination- inducing
hormones or karrikins; (iv) change in water holding capacity
and porosity of the soil. These biochar- related factors are the
reason that biochar feedstock, production HTT, and application
amount have a range of impacts on germination speed and rate.
The specific sensitivity of seeds of different plant species to
salinity, toxins, hormone- like compounds and water availability
also results in very variable results. For example, wood biochar
(HTT 620°C) at 80Mgha−1 in a pot trial inhibited germination
of tomatoes, while biochar made from paper sludge and wheat
husk (500°C) or sewage sludge (600°C), and applied at the
same rate, had no effect on lentil, tomato, cress, cucumber, and
lettuce seeds (Gascó et al., 2016). Other studies that applied a
range of woody and manure biochars at rates of 10– 40Mgha−1
1738
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JOSEPH Et al.
found positive or nil effect on germination (Das et al., 2020;
Gascó et al., 2016; Khan et al., 2014; Mete et al., 2015; Van
Zwieten, Kimber, Morris, Downie, et al., 2010). Some studies
(e.g., Uslu et al., 2020) that reported negative effects of biochar
on germination at very high rates (120Mgha−1) applied bio-
char directly to seeds in a petri dish, in the absence of soil or
other media, which is unlikely to reflect the effects of biochar in
the field environment, where charged clay minerals, microbes,
and organic compounds interact with biochar, and are likely to
modify and buffer the response. Germination rates were not af-
fected by the addition of BCF at <700kgha−1 in pot or field
trials, while seedling growth was the same or greater than with
NPK fertilizer alone (Joseph, Graber, et al., 2013; Liao et al.,
2020; Qian et al., 2014; Zheng et al., 2017). Aqueous extracts
of some biochars have been found to stimulate germination and
seedling growth (Taek– Keun et al., 2012).
Seed germination and early seedling development can be
influenced as a result of the effects of biochar on soil physi-
cal properties (Section 2.1.1). For instance, by reducing soil
bulk density and increasing soil aeration, biochar can provide
oxygen for seed germination and improve seedling growth
through lower resistance to root penetration and seedling
emergence. These effects typically increase with higher bio-
char rates (Obia et al., 2018).
Chemical impacts of biochar on soils and soil water solu-
tion can also affect seed germination and early seedling de-
velopment. For example, by raising the pH, alkaline biochars
alleviate Al and heavy metal toxicity that can reduce root
growth in acidic soils (Lauricella et al., 2021; Shetty et al.,
2020; Van Zwieten, Rose, et al., 2015). At high application
rates, biochars with high levels of soluble salts could inhibit
germination and seedling growth through osmotic stress.
Certain soluble organic compounds released from biochars
can stimulate germination and plant growth (Sun, Drosos,
et al., 2017). Kochanek et al. (2016) showed that biochars
containing karrikins, a class of water- soluble organic mol-
ecules associated with plant response to fire, can accelerate
germination and early growth of plants. These authors at-
tributed the response to signaling molecules that stimulate
plant development. The quantity of karrikins and germina-
tion response varied widely between biochars studied by
Kochanek et al. (2016). French and Iyer- Pascuzzi (2018)
found evidence that stimulation of the gibberellin pathway
contributes to the observed promotion of germination and
seedling growth by wood biochar in some tomato genotypes.
Similarly, phenols and polyphenols released from biochar
(Reynolds et al., 2018) can break seed dormancy, leading to
germination, and also promote seedling growth (Mu et al.,
2003; Stoms, 1982). Yet, some organic molecules released
can be phytotoxic, so applying biochar a few weeks in ad-
vance of sowing supports seedling growth through the de-
velopment of a beneficial rhizosphere microbiome (Jaiswal
et al., 2018).
At very high rates of application (>50Mgha−1), biochars
derived from contaminated sludges or feedstock grown in
contaminated soils can release heavy metals that inhibit ger-
mination (Das et al., 2020). Biochars contain polycyclic aro-
matic hydrocarbons (PAHs; Gascó et al., 2016; Weidemann
et al., 2018), organic pollutants formed during incomplete
combustion, that can inhibit germination at high rates.
However, PAHs in biochar are generally of little or no con-
cern for plant growth due to their strong binding by biochar,
and furthermore, their concentration is usually below regula-
tory limits if biochar is made under slow pyrolysis conditions
(Buss et al., 2015; Hale et al., 2012; Hilber et al., 2017).
At high biochar application rates in the absence of soil
(volumetrically equivalent to >40Mgha−1, in a petri dish),
free radicals from biochar inhibit germination and seedling
growth (Liao et al., 2014). However, at low biochar rates, low
levels of free radicals could be beneficial, as reactive oxygen
species can interact with plant hormones that trigger germi-
nation (Gomes & Garcia, 2013). Furthermore, free radicals
associated with biochar have been found to degrade certain
organic and inorganic pollutants (Ruan et al., 2019) which
in turn could enhance germination and seedling growth. In
addition, biochar can lower the production of reactive oxy-
gen species by plants: Natasha et al. (2021) showed that the
production of reactive oxygen species was lower, on average,
by 33% in plants grown in soils contaminated with trace ele-
ments where biochar was applied (2%– 10% w/w).
In summary, most biochars and biochar formulations do
not inhibit germination and early growth of plants in soil un-
less applied at very high rates (e.g., >40– 50Mgha−1), and
can promote germination and seedling growth at moderate
rates. The mechanisms for the positive effects largely involve
water- soluble organic compounds that stimulate germination
and seedling growth, or reactions that deactivate inhibitory
factors such as heavy metals and phytotoxic organic com-
pounds. These effects vary between biochars: low tempera-
ture biochars have a higher content of water- soluble organic
molecules that can promote germination and early growth at
low application rates; these biochars are also likely to cause
inhibition if applied at high rates. Negative effects on germi-
nation can result where high rates are applied due to release
of soluble salts or phytotoxic levels of organic compounds,
where biochar is contaminated, and where soil is absent.
Biochars with high levels of soluble mineral compounds can
also cause inhibition at high application rates.
2.2
|
Stage 2: Medium- term reactions
(1– 6months)
The effects of biochar in later periods differ from the first
stage which is dominated by dissolution of compounds from
biochar. In stage 2, plant roots intercept and interact with
|
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JOSEPH Et al.
biochar. Root hairs enter biochar pores, roots wrap around
biochar (Joseph et al., 2010; Prendergast- Miller et al., 2014),
and very small biochar particles can attach to root surfaces
(Figure 1; Chew et al., 2020). Biochar affects the abundance
of specific microorganisms especially in the rhizosphere, and
the interactions between biochar, soil, plants, and the micro-
biome affect plant growth and health (Anderson et al., 2011;
Jaiswal et al., 2015).
2.2.1
|
Physical and chemical reactions in soil
The physical and chemical properties of biochar surfaces
change significantly in Stage 2 through a range of biotic and
abiotic processes that take place in the pores exposed after
the rapid dissolution phase ends (Joseph et al., 2010). The
surface area and porosity increase (Schreiter et al., 2020),
and a fine layer of organic matter with a high concentra-
tion of C– O and C– N functional groups forms around the
external and some of the internal pore surfaces of the bio-
char and BCF. This fine layer adsorbs cations (including
heavy metals), anions, nanoparticulate minerals, and or-
ganic compounds through a range of binding mechanisms
that include cation and anion exchange, ligand exchange,
covalent bonding, complexation, chelation, precipitation,
redox, and acid– base reactions, that together result in for-
mation of organo- mineral layers (Hagemann, Joseph, et al.,
2017; Joseph, Van Zwieten, et al., 2013). These layers are
redox- active and mesoporous. Surfaces in nanopores bind
molecules more tightly than larger pores (Pignatello et al.,
2017). Some of the nutrients released from fertilizer, espe-
cially N and P, can react with the biochar pore surfaces and
organo- mineral layers (Haider et al., 2020; Hestrin et al.,
2019; Joseph et al., 2018; Kammann et al., 2015). Biochar
pores may become filled with organic matter and miner-
als, protecting organic matter from microbial decomposi-
tion (Pignatello et al., 2017) and reduces availability of
nutrients.
Microagglomerates that form on internal and external bio-
char surfaces, consisting of nanoparticulate minerals bound
with organic molecules, have a significant concentration of
– C– O, – C=O, – COOH, or – NH functional groups (Joseph
et al., 2010). Recent research indicates that many of the re-
actions described above related to biochar occur on or in the
microagglomerates.
Gases such as NH3, N2O, and CH4 produced through bi-
otic and abiotic reactions of fertilizers in soils and/or through
chemical reactions on the surfaces of the biochar can diffuse
into the nanopores (<50nm), where they can react with ox-
idants and reductants, especially if the pores contain water,
which reduces N loss and GHG emissions (Section 4.3; Chiu
& Huang, 2020; Quin et al., 2015).
2.2.2
|
Microbial responses
Meta- analyses have shown that biochar increases microbial
biomass and activities (Pokharel et al., 2020), particularly in
high- N soils (Zhang et al., 2018) and with biochars produced
at low temperature from nutrient- rich feedstocks (Li et al.,
2020). Biochars, particularly those made at low temperature
from crop residues, cause shifts in microbial community
composition, increasing the ratios of fungi to bacteria, and
gram- positive to gram- negative bacteria (Zhang et al., 2018).
The meta- analysis by Pokharel et al. (2020) identified that bi-
ochar increased microbial biomass C and the activities of the
enzymes urease, alkaline phosphatase, and dehydrogenase by
22%, 23%, 25%, and 20%, respectively, with greatest effects
in acidic fine- textured soils. This increase in enzyme activi-
ties as well as the shift in microbial community diversity and
activity (Jaiswal, Elad, et al., 2018) are directly dependent
on (i) pH increase after biochar addition, as soil acidity is
the main factor regulating microbial composition (Rousk
et al., 2010); (ii) increased aeration, and consequently, bet-
ter conditions for fungi and aerobic bacteria, as well as oxi-
dative enzymes; (iii) changes in metabolic needs due to the
prevalence of large organic compounds, and consequently,
shift in the community toward K- strategists (Cui et al., 2020),
decrease in gram- negative bacteria, shift toward saprophytic
fungi, and increase in peroxidases; and (iv) strong increase in
hydrophobic compounds in soil that favors activity of fungi
(Deng et al., 2021; Xia et al., 2020).
Li et al. (2020) noted a negative effect of high biochar
rates (>50Mg ha−1) on microbial diversity, and suggested
the following potential causes: (i) introduction of toxic com-
ponents that inhibit some species; (ii) increase in the C:N ra-
tios of SOM that limits microbial C utilization, possibly only
in the short term and only to the extent that the organic C is
metabolized; and (iii) disruption of microbial microenviron-
ments. Note also that C:N ratio does not influence microbial
metabolization of biochars (Torres- Rojas et al., 2020).
Fungi and bacteria inhabit the larger nutrient- rich pores
of biochar (>2µm) where they mine the nutrients in the bio-
char and those that have been absorbed from fertilizers. The
adsorption of root exudates, microbial metabolites, and mi-
crobial necromass increases SOM levels and thus increases
soil organic carbon (SOC; see Section 4.2). Small biochar
particles can migrate to the root surface and can alter the
abundance of specific root- associated bacteria (Chew et al.,
2020; Kolton et al., 2011).
In low P soils, arbuscular mycorrhizal fungi (AMF) invade
the pores of biochar, especially biochars with high P content
on the pore surface, which can increase plant P uptake (Gujre
et al., 2020; Solaiman et al., 2019; Vanek & Lehmann, 2015).
Blackwell et al. (2015) found that a phosphorus- enhanced
BCF increased root colonization to 75% compared with 20%
1740
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JOSEPH Et al.
in mineral fertilizer and unfertilized control and increased P
uptake efficiency.
Adsorption of microbial signaling molecules (especially
acyl- homoserine lactone) on biochar surfaces can disrupt soil
microbial communication, which could reduce the effects of
pathogens (Gao et al., 2016; Masiello et al., 2013). Biochar
can also adsorb pathogenic enzymes and toxic metabolites
exuded by soil- borne pathogens, thus reducing the concentra-
tion of virulence factors in the root zone and lowering disease
severity (Jaiswal et al., 2018).
2.2.3
|
Plant responses
Nutrient responses
Much of the N within the biochar C matrix (e.g., heterocy-
clic- N) is unavailable to plants (Clough et al., 2013; Torres-
Rojas et al., 2020), whereas most K in biochar is present in
soluble forms, released in the short term after application to
soil (Silber et al., 2010), and is readily available to plants.
Meta- analyses have found that biochar application com-
monly increases P availability, particularly when applied to
acidic or neutral soils, and for biochar produced from low
C:N feedstocks (e.g., manure, crop residues), and produced
at low temperatures (Gao et al., 2019; Glaser & Lehr, 2019).
However, P availability can be low in Ca- rich and K- poor
feedstocks such as sewage sludge (Buss et al., 2018, 2020;
Torres- Rojas et al., 2020; Wang et al., 2019) because py-
rolysis can convert plant- available organic P into inorganic
P that is less available in the short term (Buss et al., 2020;
Rose et al., 2019). The opposite has also been observed,
with pyrolysis increasing plant- available P although decreas-
ing water- extractable P (Wang et al., 2014; Zwetsloot et al.,
2015, 2016). The effect of biochar on P availability is de-
termined by microscale effects on soil pH and soil solution
composition, especially Ca content (Buss, Assavavittayanon,
et al., 2018; Buss et al., 2018). Biochar can retain nutrients,
especially N, released as fertilizers dissolve, and nutrients al-
ready present in soil, reducing loss through leaching (Haider
et al., 2020). For example, meta- analysis found that bio-
char reduces N leaching on average by 26%, though it can
increase ammonia volatilization at biochar application rates
>40Mgha−1 and with biochar pH>9 (Haider et al., 2020;
Liu, Zhang, et al., 2018). While the stimulation of micro-
bial activity by easily- mineralizable components of biochar
can reduce N availability through microbial immobilization
(Clough et al., 2013), it also accelerates the mineralization
of organic matter and nutrient cycling, and AMF root colo-
nization, which can increase N and P uptake by plants, as
discussed above (Solaiman et al., 2019) and can also improve
root growth under water stress (Mickan et al., 2016).
Adsorption of root exudates by biochar may cause disso-
lution of mineral compounds in biochar pores (Wang et al.,
2020), which can increase nutrient availability, and can re-
sult in additional adsorption sites for organic molecules
(Prendergast- Miller et al., 2014).
In flooded paddy soils, biochar and BCF particles can be
encapsulated in an organo- mineral layer (Chew et al., 2020)
on the root surface. BCF attached to the root or located in
the rhizosphere of rice grown in flooded soils was observed
to significantly alter the pH and Eh around the root, the root
membrane potential (the potential difference between the in-
side of the root and the soil), and the abundance of specific
microorganisms that increase nutrient availability (Chew
et al., 2020). Thus, when biochar is in contact with root hairs,
in the presence of microbes, it has the capacity to store and
release nutrient ions and electrons (Chew et al., 2020; Sun,
Levin, et al., 2017). The change in root membrane potential
can facilitate uptake of nutrients when required by the plant.
Chew et al. (2020) have represented these reactions as an RC
circuit (Figure 1). Biochar directly mediates electron transfer
by functioning as an electron shuttle and indirectly transfers
electrons from the valence band to the conduction band in
the Fe minerals by generating electron– hole pairs producing
reactive oxygen species (
O
⋅
−
2
, H2O2,
HO ⋅
) by Fenton and
Fenton- like reactions (Yu & Kuzyakov, 2021).
Chemolithotroph bacteria can grow on the surfaces of mi-
croagglomerates of clay and Fe nanoparticles and make S and
Fe more available to plants (Ye et al., 2017). Microbes can
form biofilms on biochar surfaces, and establish corrosion
cells that increase the solubility of metal species (e.g., insol-
uble Al2O3 to soluble Al; Joseph, Van Zwieten, et al., 2013).
Effects on heavy metal uptake
Many studies have shown that biochar can reduce uptake of
heavy metal(loid)s by plants. A meta- analysis found bio-
char addition to soils resulted in average decreases in plant
tissue concentrations of Cd, Pb, Cu, and Zn by 38%, 39%,
25%, and 17%, respectively (Chen et al., 2018). Studies
showing significant reduction in bioavailability of heavy
metals have often applied high rates of biochar, in excess
of 10Mg ha−1 (Chen et al., 2018; Wang et al., 2020). The
surface O- functional groups on biochar can immobilize
heavy metals through ion exchange, precipitation, cation
and anion metal attraction, reduction, electron shuttling,
and physisorption (Figure 2; Ahmad et al., 2014; Ding
et al., 2014; Liu, Xu, et al., 2018; Tan et al., 2015; Zheng
et al., 2020). Alkalinity from biochar (the liming effect) in-
creases pH of acid soils, increasing the negatively charged
exchange sites on clay particles, attracting cationic met-
als (Figure 1). Manure biochars commonly contain higher
Ca than plant- derived biochars, and thus can immobilize
cationic heavy metals (e.g., Cd2+ and Cu2+) through ion
exchange (Lei et al., 2019). Stable precipitates formed in
biochars with high P can immobilize Pb through the forma-
tion of β- Pb9(PO4)6, whereas higher alkalinity and calcite
|
1741
JOSEPH Et al.
in biochar facilitate the formation of insoluble hydrocerus-
site Pb3(CO3)2(OH)2 (Cao & Harris, 2010; Li et al., 2016).
Particles on the surface of biochars consisting of carbon-
coated minerals are particularly effective in reducing
bioavailability of heavy metals (Kumar & Prasad, 2018).
Incorporation into organo- mineral microagglomerates can
reduce Cr(VI) to Cr(III) through interaction with reduced
Fe, organic compounds, and free radicals (Odinga et al.,
2020), including through electron shuttling (Xu et al.,
2019), reducing their availability to plants (Kumar, Joseph,
et al., 2018; Kumar et al., 2020).
High- temperature willow biochar was found to adsorb
heavy metals from sewage sludge through both physisorption
and the mechanisms described above (Bogusz et al., 2019).
Even feedstocks that contain high contents of heavy metals
can reduce the bioavailability of some heavy metals in some
soils. For example, sewage sludge biochar decreased the bio-
accumulation of As, Cr, Co, Cu, Ni, and Pb, but increased
that of Cd and Zn in an acidic paddy soil (Khan et al., 2013).
Biochar can increase the mobility of anionic metal-
loids such as As (e.g.,
AsO3−
4
,
AsO3−
3
; Igalavithana et al.,
2017) through a decrease in positively charged sites, which
decreases the binding sites for As as soil pH increases
(Vithanage et al., 2017). Engineering biochars through add-
ing magnetite nanoparticles can increase AEC and thus ad-
sorb As (Wan et al., 2020).
Plant health
Besides the impacts of biochar on plant growth and develop-
ment, it has been observed in numerous pathosystems that bi-
ochar can elicit systemic resistance in plants against diseases
(Frenkel et al., 2017). Biochar in the growing medium can
“prime” plants (Ton & Maunch- Mani, 2003) for rapid up-
regulation of defense- related genes (Elad et al., 2010; Jaiswal
et al., 2014, 2015, 2017, 2020; Jaiswal, Elad, et al., 2018;
Kolton et al., 2017; Kumar et al., 2021; Mehari et al., 2015;
Meller Harel et al., 2012). Plants in a primed state display
faster and stronger activation of cellular defense responses,
such as earlier oxidative burst and stronger upregulation of
defense genes, upon encountering biotic stresses (Conrath
et al., 2006). This effect has been observed also for abiotic
environmental pressures such as salt, heat, cold, toxins, and
drought (Ton & Maunch- Mani, 2003).
A range of biochar– rhizosphere mechanisms are po-
tentially responsible for these in planta responses (Graber
et al., 2014), involving biochar's varied direct and indirect
influences on the soil/rhizosphere/pathogen/microbiome/
plant system. Some of these include: release of Si from bio-
char (especially straw and rice husk biochars), reported to
increase disease resistance and plant growth (Wang, Wang,
et al., 2019) by suppression of initial infection and pathogen
access to plant tissues; adsorption by biochar of extracellular
pathogenic enzymes and toxins (released by soil pathogens
FIGURE 2 Postulated mechanisms of biochar interactions with heavy metals and metalloids (adapted from Ahmad et al., 2014)
−
−
−
−
−
−
−
1742
|
JOSEPH Et al.
to dissolve and poison roots) lowering their concentrations
in the root zone (Jaiswal, Frenkel, et al., 2018); induced sys-
temic acquired resistance through upregulation of genes and
pathways associated with plant defense and growth (Jaiswal
et al., 2020); and adsorption and deactivation of plant sig-
naling molecules that induce germination of parasitic weed
seeds (Eizenberg et al., 2017).
The impact of biochar on plant disease is a function of
biochar dose and type (physical/chemical characteristics, as
discussed above; Frenkel et al., 2017; Poveda et al., 2021;
Rogovska et al., 2017). Generally, no impact is found at low
rates (<2Mg ha−1), positive impacts are seen at moderate
rates (2– 20Mgha−1), and negative impacts at relatively high
rates (>50Mgha−1). This response pattern has been observed
in studies of plant growth and disease caused by Rhizoctonia
solani in common beans (Jaiswal et al., 2015) and cucumber
(Jaiswal et al., 2014), and in other plant– soil- borne pathogen
(Graber et al., 2014) and plant– foliar pathogen (Elad et al.,
2011) systems. However, the optimal rate for disease sup-
pression does not always coincide with the optimum rate for
growth response. Rates that are beneficial for plant growth
in non- diseased systems can result in disease promotion in
pathogen- infected systems (Jaiswal et al., 2015).
Few studies have examined in planta responses to biochar
when faced with environmental pressures. Under sufficient
and drought water conditions, Chenopodium quinoa and
maize both grew significantly better in biochar treatments,
which was attributed to improved plant traits (lower pro-
line content and less negative osmotic potential) rather than
to increased root zone water content (Ahmed et al., 2018;
Kammann et al., 2011). Improved pepper plant productivity
in biochar- treated plots in a multi- year trial conducted under
extreme environmental pressures (high evaporation demand
and vapor pressure deficit, high daytime temperatures (heat
stress) at planting and low nighttime temperatures at fruiting,
brackish water irrigation) was attributed to biochar- elicited
acclimation responses in the plants (Kumar, Elad, et al.,
2018). Tests with heat stress and biochar in Arabidopsis in-
dicated early microstresses primed the plants to cope better
with subsequent acute heat stress. Early microstresses elic-
ited improved energy production and utilization mechanisms,
while the acclimation mechanism against the acute heat was
related to lower levels of reactive oxygen species. The ability
of biochar to induce an early acclimated state to basal mi-
crostresses and to prime the plant for coping with subsequent
acute stresses was postulated to explain biochar- mediated im-
provements in plant health, flowering, and growth due to fac-
tors other than nutrition, water, or soil structure (Elad et al.,
2011).
In addition to in planta responses discussed above, bio-
chars buffer pH and poise (equilibrate) Eh (Husson, 2013;
Joseph et al., 2015) which can create and maintain conditions
in the rhizosphere that support plant growth and resilience
to a range of environmental pressures, such as drought, heat,
pathogens and pollutants (Husson et al., 2018). Biochar
can rapidly transfer charge (Sun, Levin, et al., 2017; Yu &
Kuzyakov, 2021), which could also enhance plants’ capacity
to cope with oxidative stress (Husson et al., 2018).
In summary, biochar can create conditions in the rhizo-
sphere that increase nutrient supply and uptake; immobilize
or deactivate phytotoxic organic and mineral substances;
release bioactive compounds that stimulate growth and de-
velopment; promote beneficial organisms; and inhibit patho-
gens. Thus, biochar can support plant growth, health, and
resilience to disease and environmental stressors.
2.3
|
Stage 3: Long- term reactions
Several studies have examined the longer term interactions
as biochar “ages” in soil, investigating effects on bulk soil
properties and plant growth where biochar has been applied
in previous crops, or examining biochar particles extracted
from the soil. Disturbance through cultivation, exposure to
wetting– drying and freeze– thaw cycles, and ingestion by soil
fauna can lead to further fragmentation of biochar particles
and oxidation of biochar surfaces exposed through detach-
ment of microagglomerates (Wang et al., 2020).
Two studies identified the formation of porous organo-
mineral heterogeneous microagglomerates with mineral
phases consisting of Fe, Al, Si oxides, phosphates (Ca/Fe/
Al), carbonates (Ca/Mg), and chlorides (K, Na), and dimen-
sions from 1 to 50nm, bound together by organic compounds
and bonded to the biochar surface (Archanjo et al., 2017;
Rafiq et al., 2020). Simultaneous occurrence of Fe(II) and
Fe(III) present as magnetite and hematite could make N and
P more available through redox cycling of Fe (Haider et al.,
2020). This could contribute to long- term increase in P avail-
ability in response to biochar application, such as identified
in the meta- analysis of Glaser and Lehr (2019), who reported
enhancement lasting up to 5years. Aged high- temperature
wood biochar particles retain plant- available N as nitrates
and ammonium, adsorbed onto the organo- mineral micro-
agglomerates (Haider et al., 2020). The formation of micro-
agglomerates increases the surface area, CEC and AEC, but
the pore volume generally decreases compared to the fresh
biochar after multiple crop cycles, for example, Dong et al.
(2017). Rhizodeposits are protected in soil microaggregates
and Fe (oxyhydr)oxides (Jeewani et al., 2020), and a decadal
study indicates potential for this mechanism to provide long-
term stabilization of newly added plant C (Weng et al., 2017).
Biochar particles can also be protected within the soil micro-
aggregates (Figure 3).
The biochar- enriched anthropogenic Terra Preta soils
associated with pre- Columbian settlements in the Brazilian
Amazon (Steiner et al., 2009) provide evidence of very
|
1743
JOSEPH Et al.
long- term reactions of biochar in soil. Observations of Terra
Preta soils identified that a substantial fraction of the biochar
remained in particulate form, protected by Fe and Al oxides
(Glaser et al., 2000).
Colloidal aged biochar particles consisting of microag-
glomerates and fragments of the C matrix may be more mo-
bile in soil than fresh biochar (Wang, Zhang, et al., 2019).
These particles can have higher negative charge on the sur-
face compared with fresh biochar due to the higher con-
centration of C– O functional groups (Wang, Zhang, et al.,
2019), further increasing CEC and capacity to adsorb organic
molecules.
The bioavailability of heavy metals has been observed
to increase or decrease as biochar ages in soil (Wang et al.,
2020). For example, the reduction in uptake of Cd and Pb
from a highly contaminated soil was sustained over 3years
after a single application of wheat straw biochar (Bian et al.,
2014). Potentially, adding a small amount of biochar in a band
every year could ensure heavy metals remain immobilized.
In their meta- analysis, Ye et al. (2020) reported an in-
crease in crop yield over multiple years after a single bio-
char application, where fertilizer was applied. Rafiq et al.
(2020) found that moderate rates (2– 6Mgha−1) of rice husk
(high ash) biochar applied with fertilizer gave a residual
benefit for pasture yield, lasting at least 3years, associated
with enhanced microbial activity and diversity. Kumar, Elad,
et al. (2018) observed increased fruit yield and quality, and
resistance to the pathogen causing powdery mildew and the
arthropod pest broad mite, over three seasons in fertilized,
irrigated peppers after application of greenwaste and woody
biochars. Crop growth on Terra Preta soils is approximately
double that on adjacent unamended soils, providing evi-
dence that biochar can increase soil fertility over centuries
(Lehmann et al., 2003).
There is a substantial body of literature examining biochar
reactions over multiple years based on one- time application
of biochar at high rates (e.g., 20– 30Mgha−1 or 2– 3%w/w),
often in pots (e.g., Burrell et al., 2016), but there are few stud-
ies of biochar or BCF applied at low (commercially viable)
rates, as single or repeated applications. Slow release of P
from BCF and biochar can increase P- use efficiency in tropi-
cal soils over the medium- to long- term (Lustosa Filho et al.,
2020), possibly through (i) input with high P biochars such as
those made from manure or sewage sludge; and (ii) reduced
P sorption due to DOM released from biochar (Schneider &
Haderlein, 2016).
In summary, aging through interactions of biochar with soil
minerals and microbes generally leads to functionalized sur-
faces consisting of organo- mineral microagglomerates, which
can increase nutrient- holding capacity. Microagglomerates
and portions of the C matrix can detach, and colloidal- sized
particles can migrate through the soil profile. Aggregation
can protect biochar and newly added organic matter, stabiliz-
ing new C for long periods in soil. Residual effects of single
application of biochar on pH have been recorded, and some
residual yield benefits have been observed.
3
|
BIOCHAR'S ROLE IN
SUPPORTING FOOD SECURITY
Over 1700 studies published between 2010 and 2020 (Web
of Science) describe the effects of biochar on plant produc-
tion. Meta- analyses have found yield responses of annual
crops and trees of 10%– 42% and identified site and biochar
features giving greatest responses (Table 1).
Sandy soils and soils with CEC below 100mmolckg– 1
or organic C content below 20gkg−1 are most responsive
(Dai et al., 2020; Ye et al., 2020). Soil pH is consistently
identified as a key variable (Dai et al., 2020; Jeffery et al.,
2011): Responses were greatest in acidic soils, because of
the liming effect of biochar and a concomitant decline in
available Al (Van Zwieten, Rose, et al., 2015). Importantly,
Ye et al. (2020) identified that yield responses were greater
in the third year after a single application, when fertilizer
was applied with the biochar. This response most likely
reflects the physicochemical and microbial changes that
improve soil health as biochar ages (Section 2.3), rather
than a simple soil pH response. A meta- analysis by Glaser
and Lehr (2019) found availability of P increased on aver-
age by a factor of 4.6 in response to biochar application.
FIGURE 3 Fragment of biochar coated with nanoparticles that
have a high concentration of Si, Fe, Al, Ti (see Figure S4) embedded
in a soil microaggregate. Nanoparticles are the small spherical and
ovoid particles on biochar lattice. Sample of biochar removed from a
9- year field trial of greenwaste biochar (Weng et al., 2017). Mineral
nanoparticles on the biochar surface can play a key role in the
formation of microaggregates that protect biochar from decomposition.
The energy- dispersive X- ray spectroscopy (EDS) spectrum of this
image is shown in Figure S4
1744
|
JOSEPH Et al.
They noted that biochar increased P availability by a fac-
tor of 5.1 and 2.4 in acidic and neutral soils, respectively,
but it had no effect in alkaline soils or at application rates
below 10Mg ha−1 or with biochars produced at HTT <
600 °C (Glaser & Lehr, 2019). The optimal biochar dose
differed between studies (Table 1) and is dependent on the
biochar characteristics, soil properties, and the constraint
being addressed. Biochar may have no effect on yields
when low- nutrient biochars are applied without fertilizer,
or when biochar is applied to nutrient- rich soils (Ye et al.,
2020). Negative effects can result from reduction in soil
N and P availability (Nielsen et al., 2014; Prommer et al.,
2014) especially at high rates of high temperature biochars
(Kammann et al., 2015) through binding mechanisms de-
scribed in Section 2.2.
A meta- analysis of the impacts of biochar on rice produc-
tion (Awad et al., 2018) showed a net yield increase of 16%,
with greatest response at 11– 20Mgha−1 and with biochars
produced at 400– 450°C. The co- application of biochar with
N fertilizer tended to provide the greatest yield increase, sup-
porting previous evidence (Van Zwieten, Kimber, Morris,
Chan, et al., 2010) that biochar can increase fertilizer N- use
efficiency, and suggesting that biochar addition could main-
tain crop N uptake at lower doses of fertilizer N. Similarly, in
two studies using BCF, N partial factor productivity increased
by 37%– 74% (Joseph, Graber, et al., 2013; Qian et al., 2014).
Biochars are generally found to increase soil water- holding
capacity, which would enhance resilience of agricultural sys-
tems to drought, especially under climate change (Edeh et al.,
2020) and may further explain the positive effects of biochars
in sandy soils especially in arid and semiarid areas. Grass and
straw biochars increase water- holding capacity to a greater
extent than woody biochars (Burrell et al., 2016; Kroeger
et al., 2020). Meta- analyses have shown increases in plant
available water content of 33%– 45% in coarse- textured soils
and 9%– 14% in clay soils (Edeh et al., 2020; Omondi et al.,
2016; Razzaghi et al., 2020), with greatest response at 30–
70Mgha−1. Using X- ray μ- tomography, Quin et al. (2014)
observed increases in total soil porosity, connectivity of pore
space and number of fine pores across soils of different tex-
ture, explaining the results of Edeh et al. (2020) and Razzaghi
et al. (2020).
The average 27% increase in photosynthetic rate in C3
plants (but no effect on C4 plants) observed in the meta-
analysis of He et al. (2020) associated with increased stoma-
tal conductance, transpiration rate, and chlorophyll content
was attributed to the combined effects of biochar on water
availability and N nutrition.
Heavy metal pollution in arable land significantly impacts
plant growth and food safety (Luo et al., 2018) especially in
developing countries (Hou et al., 2020). Application of bio-
char to contaminated soils could reduce heavy metal bioavail-
ability via (1) direct interactions between biochar and heavy
metals, and (2) indirect interactions that immobilize heavy
metals through modification of soil properties (see Section
2.2.3), and could contribute to the yield benefits of biochar
particularly in acid soils, as soil pH is a key property gov-
erning the speciation and mobility of heavy metals. Increase
in soil CEC following biochar application can also reduce
the bioavailability of cationic heavy metals (Mohamed et al.,
2017). Biochar application can also alter soil Eh, impacting
the speciation, mobility, and bioavailability of anionic heavy
metalloids such as As (Yuan et al., 2017).
Heavy metals may be present in biochar produced
from feedstocks such as sewage sludge and treated timber.
Although the pyrolysis process concentrates most heavy met-
als, some metals such as Cd and Zn (Dong et al., 2015) and
As (Zhang et al., 2020) can be partly volatilized during py-
rolysis resulting in lower concentrations than the feedstock.
Application of biochar is a promising approach to mit-
igate heavy metal contamination; however, the remediation
efficacy depends on the type of biochar, biogeochemical
properties of soil, plant species, and the specific heavy metal
(Albert et al., 2020; Palansooriya et al., 2020). Therefore, se-
lecting the appropriate biochar type to address heavy metal
contamination, suited to the soil properties, type of plant,
and specific heavy metal, can result in effective remediation
while safeguarding food quality.
Improved understanding of the key edaphic properties that
constrain plant production and heavy metal uptake, and that
can be addressed by biochar, enables design of “bespoke bio-
chars” engineered for specific applications (Crombie et al.,
2015) to contribute to food security.
4
|
BIOCHAR'S ROLE IN CLIMATE
CHANGE MITIGATION
Biochar has been recognized as a negative emissions technol-
ogy (de Coninck et al., 2018; Cowie et al., 2020), in addition
to reducing GHG emissions from soil, as reviewed below.
Among carbon dioxide removal strategies, biochar is sug-
gested as a preferred method due to comparatively low cost
and large environmental benefits (Smith, 2016).
4.1
|
Persistent carbon in biochar
Unlike other forms of biomass that are rapidly decomposed
in soil, the majority of C in biochar has a mean residence time
in the range of hundreds and thousands of years (Schmidt
et al., 2011; Wang et al., 2016). Due to this high persistence,
biochar can contribute significantly to long- term C seques-
tration (Lehmann, 2007). Sequential additions of biochar to
soil will continue to build SOC stocks, whereas additions of
unpyrolyzed organic matter (plant litter, compost, manure)
|
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JOSEPH Et al.
will be rapidly mineralized, and will increase SOC stocks
only until an equilibrium is reached where inputs equal de-
composition rate (Figure 4).
The very slow decomposition of biochar in comparison
to unpyrolyzed biomass is attributed to its aromatic struc-
ture, which results from chemical transformations of biomass
during carbonization. Wood biochars pyrolyzed at tempera-
tures above 450– 500°C have a mean residence time of hun-
dreds to a thousand years, compared with decades for manure
biochars (Kuzyakov et al., 2014; Kuzyakov & Gavrichkova,
2009; Singh et al., 2012, 2015; Wang et al., 2016; Table 2).
Kuzyakov et al. (2009) suggested that mean residence times
calculated from incubations (Table 2), which maintain opti-
mal conditions for decomposition, are around 10 times lower
than under field conditions (Kuzyakov et al., 2009), although
Rasse et al. (2017) found a similar rate of decomposition of
Miscanthus biochar between laboratory and field conditions
over a 90 day incubation period. The kinetics of formation
of the fused aromatic C structure depend on the rate of heat-
ing, the ratio of lignin to cellulose and hemicellulose, time
at the HTT, and mineral content (Budai et al., 2014; Leng
& Huang, 2018; Rawal et al., 2016). The initial process of
drying and depolymerization is endothermic and takes place
between ambient temperatures and approximately 250°C.
This is followed by an exothermic phase where most of the
volatile gases are released, up to a temperature of approxi-
mately 350°C. The largely amorphous structure of biochars
pyrolyzed at temperatures in excess of 400– 450°C has been
found to be persistent. Further heat converts the C matrix to
a highly persistent three- dimensional nanographitic structure
at around 600°C (McDonald- Wharry et al., 2016). Minerals
present in biochar, especially Si and P, can increase per-
sistence of biochar- C (Xu et al., 2017).
Estimating potential C sequestration through the use
of biochar requires prediction of its persistence in soil.
Temperature thresholds identified in the transformation pro-
cesses can indicate persistence. Using hydrogen pyrolysis to
assess relative chemical stability, McBeath et al. (2015) esti-
mated, across a wide range of feedstocks, that <20% of the
biochar is persistent at pyrolysis temperatures <450°C, with
>80% persistent at 600– 700°C. These findings are consistent
with the structural changes observed by McDonald- Wharry
et al. (2016).
While pyrolysis temperature is a convenient measure to
obtain predictions for broad trends in persistence, and ade-
quate for national GHG inventories (Ogle et al., 2019), ma-
terial properties are a more rigorous approach to estimate
biochar persistence for project- level GHG accounting and
research applications. The elemental ratio of hydrogen to or-
ganic C expressed as H/Corg has been identified as a simple
and reliable parameter for characterizing biochar persistence
and recommendations for conservative thresholds have been
provided (Budai et al., 2013). These thresholds are being re-
fined as more data become available (Lehmann et al., 2015)
and other methods, such as spectral and thermal methods and
chemical oxidation, offer additional insights (Leng & Huang,
2018; Li & Chen, 2018).
Biochar properties are the key determinant of its per-
sistence in comparison to mineralization of unpyrolyzed bio-
mass, but edaphic and climatic factors are also influential. As
discussed in Section 2.2, the formation of microaggregates
through interaction of biochar with minerals and native SOM
FIGURE 4 Accumulation of soil organic carbon (SOC) stocks with sequential biochar additions, due to (i) the highly persistent carbon
in biochar, (ii) biochar- induced negative priming, and (iii) additional C input from plant roots through retention of rhizodeposits (Δ Root C),
compared with limited SOC stock increase with addition of unpyrolyzed organic matter. Conceptual example for a scenario where biochar is added
every 3years and decomposes at 3% per year, compared with annual additions of unpyrolyzed biomass, of which 90% decomposes each year
1746
|
JOSEPH Et al.
TABLE 2 Mean residence time (MRT) of biochars, limited to studies >150days duration
Study (experiment
period) Feedstock
Pyrolysis
temperature
(°C) Soil
MRT
(years) Study (experiment period) Feedstock
Pyrolysis
temperature
(°C) Soil
MRT
(years)
Maestrini et al. (2014) Ryegrass 450 Cambisol without
N
40 Zimmerman (2010) Oak 650 Sand 5652
(158days) Ryegrass 450 Cambisol with N 40 (365days) Grass 650 Sand 370
Santos et al. (2012) Pine 450 Graniticsoil 605 Cedar 650 Sand 15621
(180days) Pine 450 Andesitic soil 389 Bubinga 650 Sand 3937
Nguyen et al. (2014) Switchgrass 475 Typic Hapludalf 163 Sugar cane 650 Sand 7430
(189days) Switchgrass 475 Aquic Hapludult 138 Weng et al. (2015) Eucalyptus
saligna
450 Ferralsol (no plants) 484
Switchgrass 475 Lithic Dystrudept 129 (388days) E.saligna 450 Ferralsol (planted) 449
Switchgrass 475 Ultic Hapludalf 113 Wu et al. (2016) Rice straw 500 Loamy soil 857
Naisse et al. (2015)a
(222days)
Maize silage 1200 Cambisol 105 (390days) Rice straw 500 Loamy soil 2829
Maize silage (weathered) 1200 Cambisol 211 Rice straw 500 Clay loam 1896
Bai et al. (2013)a
(200days)
Miscanthus×giganteus 575 Inceptisol 28 Rice straw 500 Loamy soil 971
Miscanthus×giganteus 575 Mollisol 33 Rice straw 500 Sand clay 617
Miscanthus×giganteus 575 Inceptisol- Aquept 64 Herath et al. (2015)a Corn stover 350 Alfisol 91
Budai et al. (2016) Corncob 369 Inceptisol 252 (510days) Corn stover 350 Andisol 91
(364days) Corncob 416 Inceptisol 143 Corn stover 550 Andisol 91
Corncob 562 Inceptisol 191 Corn stover 550 Alfisol 91
Corncob 580 Inceptisol 138 Major et al. (2010) Old mango
tree
400– 600 Typic Haplustox 600
Corncob 796 Inceptisol 149 (730days)
Miscanthus×giganteus 235 Inceptisol 4 Fang et al. (2014)a E.saligna 450 Inceptisol 456
Miscanthus×giganteus 369 Inceptisol 172 (730days) E.saligna 450 Entisol 342
Miscanthus×giganteus 385 Inceptisol 165 E.saligna 450 Oxisol 342
Miscanthus×giganteus 416 Inceptisol 118 E.saligna 450 Vertisol 342
Miscanthus×giganteus 503 Inceptisol 125 E.saligna 550 Inceptisol 913
Miscanthus×giganteus 600 Inceptisol 232 E.saligna 550 Entisol 913
Miscanthus×giganteus 682 Inceptisol 123 E.saligna 550 Oxisol 685
|
1747
JOSEPH Et al.
Study (experiment
period) Feedstock
Pyrolysis
temperature
(°C) Soil
MRT
(years) Study (experiment period) Feedstock
Pyrolysis
temperature
(°C) Soil
MRT
(years)
Maestrini, Abiven,
et al. (2014)
Pine 450 Cambisol without
N
191 E.saligna 550 Vertisol 685
(365days) Pine 450 Cambisol with N 430 Zimmerman and Gao (2013)a Grass 650 Sand 104167
Zimmerman (2010) Pine 400 Sand 1280 (1173days) Oak 650 Sand 588
(365days) Oak 400 Sand 1263 Singh et al. (2012) E.saligna 400 Vertisol 294
Grass 400 Sand 793 (1829days) E.saligna
leaves
400 Vertisol 270
Cedar 400 Sand 3967 Poultry litter 400 Vertisol 129
Bubinga 400 Sand 2532 Cow manure 400 Vertisol 90
Sugar cane 400 Sand 2740 E.saligna 550 Vertisol 1616
Pine 525 Sand 2928 E.saligna
leaves
550 Vertisol 572
Oak 525 Sand 3223 Poultry litter 550 Vertisol 396
Grass 525 Sand 1218 Cow manure 550 Vertisol 313
Cedar 525 Sand 3465 Papermill
sludge
550 Vertisol 102
Sugar cane 525 Sand 1829 Kuzyakov et al. (2009)
(1181days)
Ryegrass 400 Luvisol 200
Pine 650 Sand 2284 Kuzyakov et al. (2014)
(3100days)
Ryegrass 400 Luvisol 402
aWhere not stated by the authors, we calculated MRT as the inverse of the degradation constant k2, based on a two- pool exponential model.
TABLE 2 (Continued)
1748
|
JOSEPH Et al.
can reduce the mineralization of biochar- C; thus, persistence
is likely to be greater in soils dominated by minerals that
form stable aggregates (kaolinite and sesquioxides), such as
Oxisols and Ultisols (Fang et al., 2015; Fungo et al., 2017;
Weng et al., 2017). There is some evidence that biochar per-
sistence decreases as ambient temperature increases (Fang
et al., 2017). The movement of biochar through the soil pro-
file can increase persistence in some soil types (Singh et al.,
2015).
4.2
|
Priming effects
Change in the mineralization rate of SOM induced by organic
or mineral amendments is known as “priming” (Kuzyakov
et al., 2000). Historical addition of pyrogenic organic matter
has been shown to slow SOM mineralization and enhance
native soil organic C (SOC) stocks (Borchard et al., 2014;
Downie et al., 2011; Hernandez- Soriano et al., 2016; Kerré
et al., 2016; Liang et al., 2010). The direction of priming
can be positive or negative, with an increase or decrease of
SOM mineralization, respectively. Priming effects of biochar
are reviewed in more detail in Text S3. Meta- analyses show
that biochar application commonly induces positive priming
initially (for 20days: Maestrini et al., 2015; 2years: Ding
et al., 2018), followed by negative priming, of 3.8% on aver-
age (Wang et al., 2016). Wang et al. (2016) further identi-
fied that biochar decreased SOM mineralization by 20% with
crop residue biochars, 19% with fast pyrolysis biochars, 19%
with low temperature biochars (200– 375°C), and 12% with
low biochar application rates (0.1%– 1%w/w) but increased
SOM mineralization by 21% in sandy soils. Ding et al. (2014)
found that the magnitude of negative priming increased with
increasing pyrolysis temperature, time following biochar ap-
plication, and soil clay content >50%, but decreased with an
increasing C:N ratio of soil.
Mechanisms for biochar- induced positive priming in-
clude direct effects from: (1) greater microbial activity
and enzyme production fueled by the addition of the eas-
ily mineralizable C from biochar (Luo et al., 2013; Singh
& Cowie, 2014; Section 2.1), and (2) microbial nutrient
mining (e.g., N and P); and indirect effects such as (1)
amelioration of acidity by biochar that promotes microbial
activities (Luo et al., 2011), (2) amelioration of nutrient
constraints (Mukherjee & Zimmerman, 2013), (3) en-
hanced microbial habitat (Luo et al., 2013; Pokharel et al.,
2020) and soil faunal activity, and (4) much better aeration
because of increased size and stability of macroaggregates
and lower soil bulk density, all leading to increased micro-
bial activities.
Biochar can cause negative priming directly by (1) sub-
strate switching where the easily mineralizable C from
biochar may be preferentially consumed by microbes
to temporarily replace the use of SOC (DeCiucies et al.,
2018; Kuzyakov et al., 2000) and (2) a dilution effect of
substrates where added biochar temporarily reduces the
mineralization of the more readily mineralizable C in soil
(Whitman et al., 2014) and indirectly from (3) the sorp-
tion of organic compounds by biochar (DeCiucies et al.,
2018; Kasozi et al., 2010), (4) improved organo- mineral
protection and stable aggregation slowing down the min-
eralization of SOC within the organo- mineral complexes
(Fang et al., 2018; Weng et al., 2017, 2018), and (5) in-
hibition of microbial activity by polyaromatic toxic com-
pounds (Zhang et al., 2018). Biochar amendments reduced
the activities of soil enzymes associated with C cycling by
6% (Zhang et al., 2019), improved C- use efficiency (Liu,
Zhu, et al., 2019, 2020), increased soil microbial biomass
(Li et al., 2020), and lowered the metabolic quotient by
12%– 21% (i.e., respiration rate CO2- C per unit of micro-
bial biomass C) compared with the unamended soils (Zhou,
Zhang, et al., 2017), the latter attributed to improved mi-
crobial habitats and alleviation of environmental stresses
including acid soil constraints. Negative priming is found
to result mainly from substrate switching (Ventura et al.,
2019) and dilution (DeCiucies et al., 2018) in the short
term, with adsorption being more important after several
weeks.
Biochar can affect new additions to soil of plant- derived
C, and these rhizodeposits can also prime and act as a source
of SOC. In a subtropical pasture on a Rhodic Ferralsol, a
13C- depleted hardwood biochar (450°C) initiated positive
priming up to 0.15MgCha−1 over 62days, switching to
negative priming after 188days in the presence of plants
(Weng et al., 2015). Biochar builds SOC through soil ag-
gregation processes that stabilize new C (i.e., rhizodepos-
its), by 6%– 16% (Ventura et al., 2019; Weng et al., 2015,
2017), as well as by reducing priming caused by plant C
input (Whitman et al., 2014). In a 6- year field experiment
where woody biochar was applied to corn and bioenergy
crops, SOC stocks increased by 14MgCha−1, twice the
quantity of C added in the biochar, as a result of negative
priming (Blanco- Canqui et al., 2020). Figure 4 illustrates
how biochar application can lead to accumulation of SOC
stocks through biochar- induced negative priming and en-
hanced retention of rhizodeposits.
4.3
|
Effect on GHG emissions
The complex soil microbial communities that produce
and consume N2O and CH4 in soil and the interrelated
biotic and abiotic processes that take place, make pre-
dicting GHG emissions from soil extremely challenging.
Microbiological N transformations are the main source
of N2O emissions from soil, with autotrophic nitrification
|
1749
JOSEPH Et al.
and heterotrophic denitrification being the main N2O for-
mation pathways. Biochar can lower denitrification (the
reduction of
NO −
3
to N2) by: (i) facilitating the last step
of denitrification (the transformation of N2O to N2), and
(ii) decreasing total denitrification activity (Cayuela et al.,
2013; Weldon et al., 2019). Biochar can facilitate the re-
duction of N2O to N2 via: (i) increasing pH in acid soils
(Obia et al., 2015) thus enhancing the nosZ gene (Harter,
Guzman- Bustamante, et al., 2016); (ii) changing the rela-
tive abundance and composition of N2O- reducing micro-
bial communities (Harter, Weigold, et al., 2016); and (iii)
facilitating extracellular electron exchange (Chen et al.,
2014) or directly donating electrons to denitrifying bac-
teria (Pascual, Sánchez- Monedero, Cayuela, et al., 2020).
The decrease in denitrification may result from decrease
in availability of NO3− and bioavailable C substrate
(Fiorentino et al., 2019; Hagemann, Kammann, et al., 2017;
Heaney et al., 2020). Abiotic processes, in particular with
biochars containing high Fe and Mn content (see Section
2.2.1), can directly catalyze the reduction of N2O to N2. It
has also been shown that N2O can be transformed to NH3,
pyridine, or pyrrole compounds on biochar surfaces, thus
decreasing N2O emissions (Quin et al., 2015).
Several meta- analyses have synthesized the results of
studies on effects of biochar on soil GHG emissions, and
sought to explain the differences between individual studies.
Although there are gaps in process understanding, and identi-
fication of best management practices, there is solid evidence
that biochar can mitigate soil N2O and CH4 emissions from
soil, at least in the short and medium term (Borchard et al.,
2019; Cayuela et al., 2014, 2015; Fan et al., 2017; Jeffery
et al., 2016; Liu, Liu, et al., 2019; Liu, Zhang, et al., 2018;
Verhoeven et al., 2017).
Early meta- analyses on N2O emissions showed very high
mitigation (around 50% reductions) of N2O with biochar
(Cayuela et al., 2014, 2015). These studies included labora-
tory experiments performed under controlled conditions, and
with very high biochar application rates (>100Mgha−1). A
direct correlation between application rate and N2O decrease
was found (Cayuela et al., 2014), with lower N2O mitiga-
tion (average 27%) under more realistic rates equivalent to
10– 20Mgha−1. Most experiments included in these meta-
analyses were carried out under high moisture conditions
favoring denitrification, where biochar is most effective in
decreasing N2O emissions (Cayuela et al., 2013; Weldon
et al., 2019).
Later meta- analyses including a larger number of field
studies and more realistic biochar application rates found
lower average reductions, of 12% (Verhoeven et al., 2017)
considering only field studies, and 38% (Borchard et al.,
2019) including laboratory and field studies. This contrasts
sharply with other (unpyrolyzed) organic amendments. For
example, a meta- analysis on manure application to soil found
an average increase of 33% in N2O emissions compared to
synthetic fertilizer (Zhou, Zhu, et al., 2017). Even high C:N
amendments that tend to immobilize N in soil have been
found to increase N2O emissions. For instance, Xia et al.
(2018) found an average increase of 22% in N2O emissions
when straw was applied. Therefore, although the averaged
numbers differ between meta- analyses depending on the cri-
teria for the inclusion of studies and the methodology used,
there is strong evidence that biochar amendment reduces (on
average) direct N2O emissions from soil particularly when
compared to other organic amendments.
Biochars produced by slow pyrolysis, with high degree
of carbonization, high pH, and high surface area, are most
effective in suppressing N2O emissions (Borchard et al.,
2019; Cayuela et al., 2015; Weldon et al., 2019). A dose of
10– 20Mgha−1 has been found to significantly reduce N2O
emissions (Borchard et al., 2019; Cayuela et al., 2014). The
effect of biochar might diminish with time, as biochar ages
in soil (Borchard et al., 2019; Fungo et al., 2017; Liu, Zhang,
et al., 2018). Nevertheless, the mitigation provided initially
can be substantial, and repeated applications may maintain
the mitigation benefit.
The impact of biochar on CH4 fluxes has been widely
evaluated in paddy and non- flooded soils. Whereas non-
flooded soils mostly act as a sink of atmospheric CH4,
paddy soils can be a significant source of CH4. Several meta-
analyses found that, on average, biochar mitigates CH4 emis-
sions from flooded soils, particularly from acidic soils, but
decreases the CH4 sink of non- flooded soils (Jeffery et al.,
2016). Ji et al. (2018) cautioned that the co- application of
biochar with nitrogen fertilizers substantially decreased the
effectiveness of biochar in reducing soil CH4 emissions from
paddies, however, their meta- analysis also showed that the
biochar- induced decrease in CH4 uptake by non- flooded soils
was lessened when N fertilizer was also applied. Further, a
recent study demonstrates the relevance of biochar proper-
ties to the effect on soil CH4 uptake rates: biochars with high
electrical conductivity and ash concentrations decreased CH4
sink capacity whereas biochars from woody materials pyro-
lyzed at high temperatures and with high pore area increased
soil CH4 uptake rates (Pascual et al., 2020). Qian et al. (2014)
found a decrease in N2O and CH4 emissions from paddy soil
when a range of biochar- based BCFs was compared with
NPK fertilizers.
4.3.1
|
GHG intensity and yield-
scaled emissions
To avoid overlooking potential trade- offs with crop yields,
studies report GHG intensity (GHG per unit crop yield)
1750
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JOSEPH Et al.
(Mosier et al., 2006) or yield- scaled emissions for N2O
(i.e., N2O emissions in relation to N uptake of the above-
ground crop) (Van Groenigen et al., 2010). Analyses of
specific cropping systems show a decrease in GHG in-
tensity with biochar application in vegetable fields (Fan
et al., 2017) and in wheat– rice rotation systems (Wu et al.,
2019). One of the first studies summarizing results on
yield- scaled N2O emissions was performed by Verhoeven
et al. (2017) who found that biochar decreased yield-
scaled N2O emissions across the majority of the studied
cropping systems, although a meta- analysis could not be
carried out due to the low number of field studies and ex-
cessively high variance between studies. Later, Liu, Mao,
et al. (2019) were able to incorporate a larger number of
studies and showed an overall reduction of GHG intensity
by 29% after biochar amendment, with higher reductions
in non- flooded soils (−41%) compared to paddy fields
(−17%). A meta- analysis focusing on vegetable fields in
China also found that biochar application decreased yield-
scaled N2O emissions by an average of 35% (Gu et al.,
2020).
4.3.2
|
Potential trade- offs between C
sequestration and non- CO2 GHG emissions
In order to evaluate the full net GHG balance of biochar in
soil, the fluxes of CH4 and N2O and the changes in SOC
stocks need to be jointly assessed. Usually, CH4 and N2O
emissions are expressed in CO2- equivalents using 100-
year global warming potential. In non- flooded soils, the
relationship between SOC changes and N2O emissions
usually regulates the net GHG emission, since agricultural
soils are often weak CH4 sinks. One of the greatest dif-
ficulties for the comprehensive analysis of the balance be-
tween C sequestration and N2O emission lies in the need
for long- term studies to measure changes in SOC reserves
(Smith et al., 2020) and the laborious nature of direct
measurements of N2O, which makes long- term N2O stud-
ies (>10years) very rare.
An increase in SOC is often associated with higher N2O
emissions, which could counteract the mitigation benefits de-
rived from C sequestration (Davies et al., 2020). However, it
is precisely in these trade- offs where biochar might have the
greatest advantage compared to other soil amendments and
other SOC sequestration strategies. Although a comprehen-
sive meta- analysis on these trade- offs has not been published
yet, results from separate meta- analyses on C sequestration
(Bai et al., 2019) and N2O emissions (Borchard et al., 2019;
Liu, Liu, et al., 2019) point to a strong synergy between C
sequestration and mitigation of N2O emissions with biochar,
which is much less evident for other SOC sequestration strat-
egies (Guenet et al., 2021).
5
|
BIOCHAR'S ROLE IN THE
CIRCULAR ECONOMY
The circular economy concept aims to conserve resources,
and minimize inputs and waste. Biochar can support the de-
velopment of a circular economy at regional and farm scale
by improving nutrient recovery and nutrient use efficiency.
The economic case for biochar production is strongest for
biochar made from residue materials, especially when the
residues contain high concentrations of nutrients, such as
animal manures and sewage sludge. Concerns that these
feedstocks may contain contaminants restrict their beneficial
reuse. Fortunately, most organic contaminants are destroyed
with high efficiency during pyrolysis, by thermal degrada-
tion and volatilization followed by destruction during vapor
combustion. This has been shown for PAHs (Zielińska &
Oleszczuk, 2015), polychlorinated biphenyls (Bridle et al.,
1990), per- and polyfluoroalkyl substances (PFAS; Kundu
et al., 2021), microplastics (Ni et al., 2020), antimicrobials
(Ross et al., 2016), antibiotics (Tian et al., 2019), antibiotic
resistance genes (Kimbell et al., 2018), and hormones (estro-
gen; Hoffman et al., 2016).
While incineration destroys organic contaminants with
similar efficiency to pyrolysis (Baukal et al., 1994), unlike
incineration, pyrolysis retains a large portion of the feed-
stock C (typically around 50%), and most nutrients, in the
biochar. In addition, pyrolysis gases can be captured for use
as a renewable energy product (see Text S4). Of the main
plant nutrients, P and K are fully retained in biochar at typi-
cal pyrolysis temperatures (300°– 700°) (Bridle & Pritchard,
2004; Buss et al., 2016). Nonetheless, 50%– 80% of N can be
lost (Hossain et al., 2011; Ye et al., 2020; Yuan et al., 2018)
depending on the N content of the feedstock (Torres- Rojas
et al., 2020), with greater loss at high pyrolysis temperature.
A meta- analysis found N, P, and K concentrations in biochars
of 1.0%, 0.4%, and 1.9% (wood- derived biochars), 1.5%,
0.8%, and 4.1% (crop residue biochars) and 2.4%, 2.6%, and
2.5% (manure/sewage sludge biochars), respectively (Ippolito
et al., 2020).
Notably, some sewage sludge biochars contain as much
as 6%– 20% total P (Faria et al., 2018; Roberts et al., 2017;
Shepherd et al., 2016; Zhang et al., 2015). However, only
a fraction of the total nutrients in biochar is available for
plant uptake (in the short- medium term), in the order
K>P>N. A meta- analysis found that, on average, the fol-
lowing percentages of the N, P and K present in biochar
were bioavailable: 0.5%, 3%, and 9% (wood- derived bio-
char), 0.4%, 6%, and 22% (crop residue biochar) and 5%,
5%, and 17% (manure/sewage sludge biochar), respectively
(Ippolito et al., 2020).
Biochar P availability can be increased by selecting low
Ca feedstocks or doping feedstock with K, leading to pref-
erential binding of P with K instead of Ca, Mg, Fe, or Al,
|
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JOSEPH Et al.
forming highly soluble salts (Buss et al., 2020). Biochars can
be optimized to sorb P or N from wastewater and hence be
loaded with extra nutrients that are accessible to plants (Mood
et al., 2020; Shang et al., 2018; Shepherd et al., 2016, 2017),
reducing wastewater P and N concentrations, preventing eu-
trophication, and returning nutrients to agricultural land.
Controlled release biochar- fertilizer combinations can
be produced from low- nutrient biomass mixed with mineral
or organic nutrients before pyrolysis and/or organic nutri-
ents after pyrolysis, or by composting to enrich with nutri-
ents (Buss et al., 2019, 2020; Dong et al., 2019; Hagemann,
Joseph, et al., 2017; Schmidt et al., 2015) and these can be
effective at low application rates when applied in a band near
the seed/plant (Qian et al., 2014; Schmidt et al., 2015; Yao
et al., 2015; Zheng et al., 2017).
The use of biochar in composting of organic residues such
as manures can reduce N losses through volatilization and
leaching, reduce GHG emissions, increase C persistence,
and reduce availability of heavy metals (Agyarko- Mintah
et al., 2017; Akdeniz, 2019; Oldfield et al., 2018; Sanchez-
Monedero et al., 2018).
Biochars, including BCFs and biochar used as a compost
additive, thus improve nutrient recovery from organic resi-
dues, facilitate use of residues in soil amendment, and re-
duce environmental impacts of waste management. Biochar
systems (Figure 5) thereby contribute to building a circular
economy.
6
|
CONCLUSION AND
RECOMMENDATIONS
Soil and plant responses to addition of biochar can be nega-
tive, positive, or neutral, depending on many variables, in-
cluding feedstock and pyrolysis temperature, application
rate and method, and application context (crop, soil type,
and environmental and biological stresses). Considering
the heterogeneous nature of biochars and the complexity of
the physical, biochemical, and microbiological processes
underpinning the effects of biochars, reviewed above, it is
not surprising that studies report a wide range of responses
to biochar application. Results are also strongly influenced
by experimental design aspects; studies that do not include
plants, or are undertaken in soil- less media, or based on pot
trials cannot readily be extrapolated to field situations.
Scientific understanding of the biochar– soil– plant pro-
cesses and interactions has evolved over the last decade,
providing the basis to interpret the divergent results in the lit-
erature and identify optimal uses of biochars. The following
encapsulates current knowledge, as reviewed in this paper.
Biochar catalyzes microbial and abiotic processes in the
rhizosphere, decreasing the activation energy for biotic and
abiotic reactions, which can increase nutrient mineralization
and facilitate nutrient uptake by plants. Higher microbial ac-
tivities lead to accelerated turnover of organic matter which
enhances nutrient supply. Biochar reduces the availability of
heavy metals, increases plant resistance to disease, and im-
proves resilience to environmental stressors. The microscale
processes on the biochar surface and in the rhizosphere me-
diate the macro responses of plants to biochar. The catalytic
ability of biochar changes as it ages in soil through oxidation
and interactions with minerals, microbes, soil fauna, and or-
ganic matter.
Significant yield increases occur where site- specific soil
constraints, nutrient and water limitations are addressed by
appropriate biochar formulations applied at an optimal appli-
cation rate. Meta- analyses of crop responses to biochar show
average yield increases of 10%– 42%, with greatest responses
in acidic and sandy soils where the biochar has been applied
with organic and/or mineral fertilizers. On average, biochars
FIGURE 5 Biochar systems utilize organic residues, including forest, crop, and horticultural residues, to produce biochar that is used as a soil
amendment directly, and indirectly via feeding to livestock. Pyrolysis gases and process heat, co- products of biochar production, can be used to
supply renewable energy
1752
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JOSEPH Et al.
increase P availability by a factor of 4.6, decrease plant tis-
sue concentration of heavy metals by 17%– 39%, build SOC
through negative priming by 3.8% (range −21% to +20%),
and reduce non- CO2 GHG emissions from soil by 12%– 50%.
To enable widespread adoption, biochar needs to be read-
ily integrated with farming operations, and be economically
viable. Formulations that combine biochar with mineral and/
or organic fertilizers and minerals are likely to have high nu-
trient use efficiency and be the most cost- effective. Such for-
mulations are the major focus of commercialization, but they
have received limited attention in research studies, and very
few field trials have been undertaken.
Knowledge gaps remain regarding biochar– soil– plant
interactions in the field over the longer term, including lon-
gevity of yield response and reduction of N2O emissions;
the direction, magnitude, and duration of organic matter
priming; and long- term effects of repeated applications.
Research is needed on processes that influence the capture
and release of heavy metals in the long term to determine
optimum scheduling of re- application of biochar. Further
research on the effects of biochar properties on root mem-
brane potential and microbial nutrient cycling will inform
the development of optimal formulations to increase nutri-
ent uptake efficiency.
We recommend that guidelines on selecting and pro-
ducing biochar formulations to meet specific soil and en-
vironmental constraints and increase farm profitability be
developed, based on the findings of this review. Biochars can
be tailored for specific applications through feedstock selec-
tion; by modifying process conditions; through pre- or post-
production treatments to adjust pH, increase nutrient level
and availability, carbon persistence and adsorptive proper-
ties; or co- application with organic or mineral fertilizers. Use
of biochar in waste management, such as co- composting of
animal manures and pyrolysis of sewage sludge, can capture
nutrients and reduce GHG emissions.
This review presents strong evidence that biochar can
contribute to climate change mitigation through carbon se-
questration and reduction in soil GHG emissions, and that
significant benefits to plant production are possible, par-
ticularly where site- specific soil constraints and nutrient
and water limitations are addressed by appropriate biochar
and fertilizer applications. Biochar has the greatest poten-
tial to increase crop yields in low- nutrient, high P- fixing
acidic soils, common in the tropics and humid subtropics,
and in sandy soils, particularly in dryland regions that are
likely to be increasingly affected by drought under climate
change. Biochar can also mitigate heavy metal pollution,
that impacts food production and food safety in many de-
veloping countries, and enhance resource use efficiency.
Thus, biochar can play a key role in addressing climate
change and supporting global food security and the circu-
lar economy.
ACKNOWLEDGMENTS
We thank Bhupinderpal Singh, Aaron Simmons, and two
anonymous referees for helpful comments on the manuscript.
The graphic design, by Anders Claassens, was funded by
La Trobe University’s Research Focus Area Collaboration
Ready grant in Securing Food, Water and the Environment
(LTU SFWE RFA 2000004349). Y.K. acknowledges sup-
port of the Program of Competitive Growth of Kazan Federal
University and the “RUDN University program 5- 100.”
CONFLICT OF INTEREST
The authors receive research funding from a range of gov-
ernment and industry sources. SJ and ALC are members of
the Australia New Zealand Biochar Industry Group Advisory
Board.
DATA AVAILABILITY STATEMENT
Data sharing is not applicable to this article as no new data
were created in this study.
ORCID
Stephen Joseph https://orcid.org/0000-0002-8933-8010
Annette L. Cowie https://orcid.org/0000-0002-3858-959X
Lukas Van Zwieten https://orcid.
org/0000-0002-8832-360X
Nanthi Bolan https://orcid.org/0000-0003-2056-1692
Alice Budai https://orcid.org/0000-0002-6675-4548
Wolfram Buss https://orcid.org/0000-0002-9653-0895
Maria Luz Cayuela https://orcid.
org/0000-0003-0929-4204
Ellen R. Graber https://orcid.org/0000-0002-4217-3204
James A. Ippolito https://orcid.org/0000-0001-8077-0088
Yakov Kuzyakov https://orcid.org/0000-0002-9863-8461
Yu Luo https://orcid.org/0000-0002-3834-498X
Yong Sik Ok https://orcid.org/0000-0003-3401-0912
Kumuduni N. Palansooriya https://orcid.
org/0000-0001-5907-3827
Zhe (Han) Weng https://orcid.org/0000-0002-9567-095X
Johannes Lehmann https://orcid.
org/0000-0002-4701-2936
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SUPPORTING INFORMATION
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the Supporting Information section.
How to cite this article: Joseph, S., Cowie, A. L., Van
Zwieten, L., Bolan, N., Budai, A., Buss, W., Cayuela,
M. L., Graber, E. R., Ippolito, J. A., Kuzyakov, Y., Luo,
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biochar works, and when it doesn't: A review of
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