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Biochar as a tool to reduce the agricultural greenhouse-gas burden – knowns, unknowns and future research needs

Authors:
  • Ithaka Insitute gGmbH, Germany and Agroscope, Zurich Reckenholz, Switzerland
  • German Agricultural Society e.V.

Abstract and Figures

Agriculture and land use change has significantly increased atmospheric emissions of the non-CO2 green-house gases (GHG) nitrous oxide (N2O) and methane (CH4). Since human nutritional and bioenergy needs continue to increase, at a shrinking global land area for production, novel land management strategies are required that reduce the GHG footprint per unit of yield. Here we review the potential of biochar to reduce N2O and CH4 emissions from agricultural practices including potential mechanisms behind observed effects. Furthermore, we investigate alternative uses of biochar in agricultural land management that may significantly reduce the GHG-emissions-per-unit-of-product footprint, such as (i) pyrolysis of manures as hygienic alternative to direct soil application, (ii) using biochar as fertilizer carrier matrix for underfoot fertilization, biochar use (iii) as composting additive or (iv) as feed additive in animal husbandry or for manure treatment. We conclude that the largest future research needs lay in conducting life-cycle GHG assessments when using biochar as an on-farm management tool for nutrient-rich biomass waste streams.
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Journal of Environmental Engineering and Landscape
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ISSN: 1648-6897 (Print) 1822-4199 (Online) Journal homepage: http://www.tandfonline.com/loi/teel20
Biochar as a tool to reduce the agricultural
greenhouse-gas burden – knowns, unknowns and
future research needs
Claudia Kammann, Jim Ippolito, Nikolas Hagemann, Nils Borchard, Maria
Luz Cayuela, José M. Estavillo, Teresa Fuertes-Mendizabal, Simon Jeffery,
Jürgen Kern , Jeff Novak, Daniel Rasse, Sanna Saarnio, Hans-Peter Schmidt,
Kurt Spokas & Nicole Wrage-Mönnig
To cite this article: Claudia Kammann, Jim Ippolito, Nikolas Hagemann, Nils Borchard, Maria
Luz Cayuela, José M. Estavillo, Teresa Fuertes-Mendizabal, Simon Jeffery, Jürgen Kern , Jeff
Novak, Daniel Rasse, Sanna Saarnio, Hans-Peter Schmidt, Kurt Spokas & Nicole Wrage-Mönnig
(2017) Biochar as a tool to reduce the agricultural greenhouse-gas burden – knowns, unknowns
and future research needs, Journal of Environmental Engineering and Landscape Management,
25:2, 114-139, DOI: 10.3846/16486897.2017.1319375
To link to this article: http://dx.doi.org/10.3846/16486897.2017.1319375
© 2017 The Author(s) Published by VGTU
Press and Informa UK Limited, [trading as
Taylor & Francis Group].
Published online: 28 Jun 2017.
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MICROBIAL COMMUNITY CHANGES IN TNT SPIKED SOIL BIOREMEDIATION
TRIAL USING BIOSTIMULATION, PHYTOREMEDIATION AND
BIOAUGMENTATION
Hiie No
˜lvak
1
, Jaak Truu
2
, Baiba Limane
3
, Marika Truu
4
,
Guntis Cepurnieks
5
, Vadims Bartkevicˇs
6
, Jaanis Juhanson
7
, Olga Muter
8
1, 7
Institute of Molecular and Cell Biology, Faculty of Science and Technology, University of Tartu,
23 Riia str., 51010 Tartu, Estonia
1, 2, 4
Institute of Ecology and Earth Sciences, Faculty of Science and Technology, University of Tartu,
46 Vanemuise str., 51014 Tartu, Estonia
3, 8
Institute of Microbiology and Biotechnology, University of Latvia, 4 Kronvalda blvd.,
LV-1586 Riga, Latvia
4, 5, 6
Institute of Food Safety, Animal Health and Environment (BIOR), 3 Lejupes str.,
LV-1076 Riga, Latvia
Submitted 6 Mar. 2012; accepted 14 Aug. 2012
Abstract. Trinitrotoluene (TNT), a commonly used explosive for military and industrial applications, can cause
serious environmental pollution. 28-day laboratory pot experiment was carried out applying bioaugmentation using
laboratory selected bacterial strains as inoculum, biostimulation with molasses and cabbage leaf extract, and
phytoremediation using rye and blue fenugreek to study the effect of these treatments on TNT removal and changes
in soil microbial community responsible for contaminant degradation. Chemical analyses revealed significant
decreases in TNT concentrations, including reduction of some of the TNT to its amino derivates during the 28-day
tests. The combination of bioaugmentation-biostimulation approach coupled with rye cultivation had the most
profound effect on TNT degradation. Although plants enhanced the total microbial community abundance, blue
fenugreek cultivation did not significantly affect the TNT degradation rate. The results from molecular analyses
suggested the survival and elevation of the introduced bacterial strains throughout the experiment.
Keywords: TNT, bioaugmentation, biostimulation, phytoremediation, microbial community.
Reference to this paper should be made as follows: No
˜lvak, H.; Truu, J.; Limane, B.; Truu, M.; Cepurnieks, G.;
Bartkevicˇs, V.; Juhanson, J.; Muter, O. 2013. Microbial community changes in TNT spiked soil bioremediation trial
using biostimulation, phytoremediation and bioaugmentation, Journal of Environmental Engineering and Landscape
Management 21(3): 153162. http://dx.doi.org/10.3846/16486897.2012.721784
Introduction
The nitroaromatic explosive, 2,4,6-trinitrotoluene (TNT),
has been extensively used for over 100 years, and this
persistent toxic organic compound has resulted in soil
contamination and environmental problems at many
former explosives and ammunition plants, as well as
military areas (Stenuit, Agathos 2010). TNT has been
reported to have mutagenic and carcinogenic potential
in studies with several organisms, including bacteria
(Lachance et al. 1999), which has led environmental
agencies to declare a high priority for its removal from
soils (van Dillewijn et al. 2007).
Both bacteria and fungi have been shown to
possess the capacity to degrade TNT (Kalderis et al.
2011). Bacteria may degrade TNT under aerobic or
anaerobic conditions directly (TNT is source of carbon
and/or nitrogen) or via co-metabolism where addi-
tional substrates are needed (Rylott et al. 2011). Fungi
degrade TNT via the actions of nonspecific extracel-
lular enzymes and for production of these enzymes
growth substrates (cellulose, lignin) are needed. Con-
trary to bioremediation technologies using bacteria or
bioaugmentation, fungal bioremediation requires
an ex situ approach instead of in situ treatment (i.e.
soil is excavated, homogenised and supplemented
with nutrients) (Baldrian 2008). This limits applicabil-
ity of bioremediation of TNT by fungiin situ at a field
scale.
Corresponding author: Jaak Truu
E-mail: jaak.truu@ut.ee
JOURNAL OF ENVIRONMENTAL ENGINEERING AND LANDSCAPE MANAGEMENT
ISSN 1648-6897 print/ISSN 1822-4199 online
2013 Volume 21(3): 153162
doi:10.3846/16486897.2012.721784
Copyright ª2013 Vilnius Gediminas Technical University (VGTU) Press
www.tandfonline.com/teel
© 2017 The Author(s) Published by VGTU Press and Informa UK Limited,
[trading as Taylor & Francis Group].
This is an Open Access article distributed under the terms of the Creative Commons Attribution-Non-
commercial-No Derivatives Licence (http://creativecommons.org/licenses/by-nc-nd/4.0/), which permits
non-commercial re-use, distribution, and reproduction in any medium, provided the original work is
properly cited, and is not altered, transformed, or built upon in any way.
The special issue on
Biochar as an Option for Sustainable Resource Management
Corresponding author: Claudia Kammann
E-mail: claudia.kammann@hs-gm.de
JOURNAL OF ENVIRONMENTAL ENGINEERING AND LANDSCAPE MANAGEMENT
ISSN 1648–6897 / eISSN 1822-4199
2017 Volume 25(02): 114–139
https://doi.org/10.3846/16486897.2017.1319375
Review article
BIOCHAR AS A TOOL TO REDUCE THE AGRICULTURAL GREENHOUSE
GAS BURDEN  KNOWNS, UNKNOWNS AND FUTURE RESEARCH NEEDS
Claudia KAMMANNa, Jim IPPOLITOb, Nikolas HAGEMANNc, Nils BORCHARDd, Maria Luz CAYUELAe,
José M. ESTAVILLOf, Teresa FUERTES-MENDIZABALf, Simon JEFFERYg, Jürgen KERNh, Je NOVAKi,
Daniel RASSEj, Sanna SAARNIOk, Hans-Peter SCHMIDTl, Kurt SPOKASm, Nicole WRAGE-MÖNNIGn
aDepartment of Soil Science and Plant Nutrition, WG Climate Change Research for Special Crops,
Hochschule Geisenheim University, Von-Lade Str. 1, 65366 Geisenheim, Germany
bDepartment of Soil and Crop Sciences, Colorado State University, Fort Collins, CO 80523-1170, USA
cGeomicrobiology, Center for Applied Geosciences, University Tübingen, Hölderlinstr. 12, 72074 Tübingen, Germany
dCenter for International Forestry Research, Jalan CIFOR, Situ Gede, Sindang Barang, Bogor 16115, Indonesia
eDepartment of Soil and Water Conservation and Waste Management, CEBAS-CSIC,
Campus Universitario de Espinardo, 30100 Murcia, Spain
fDepartment of Plant Biology and Ecology, University of the Basque Country (UPV/EHU),
Apdo. 644, E-48080 Bilbao, Spain
gCrop and Environment Sciences Department, Harper Adams University, Newport, Shropshire,
TF10 8NB, United Kingdom
hDepartment of Bioengineering, Leibniz Institute for Agricultural Engineering and Bioeconomy,
Max-Eyth-Allee 100, 14469 Potsdam, Germany
iUnited States Department of Agriculture, Agricultural Research Service (USDA-ARS),
Water and Plant Conservation Research, 2611 W Lucas street, Florence, South Carolina 295011242, USA
jDepartment of Soil Quality and Climate Change, Norwegian Institute of Bioeconomy Research,
Høgskoleveien 7, 1430 Aas, Norway
kDepartment of Environmental and Biological Sciences, University of Eastern Finland,
P.O. Box 111, 80101 Joensuu, Finland
lIthaka Institute for Carbon Strategies, Ancienne Eglise 9, CH-1974 Arbaz, Switzerland
mUnited States Department of Agriculture, Agricultural Research Service (USDA-ARS) and University of
Minnesota, Department of Soil, Water, and Climate, 1529 Gortner Ave., St. Paul, MN 55108, USA
nGrassland and Fodder Sciences, Faculty of Agriculture and the Environment, University of Rostock,
Justus-von-Liebig-Weg 6, 18051 Rostock, Germany
Submitted 7 Jan. 2017; accepted 11 Apr. 2017
Abstract. Agriculture and land use change has signicantly increased atmospheric emissions of the non-CO2 green-
house gases (GHG) nitrous oxide (N2O) and methane (CH4). Since human nutritional and bioenergy needs continue
to increase, at a shrinking global land area for production, novel land management strategies are required that reduce
the GHG footprint per unit of yield. Here we review the potential of biochar to reduce N2O and CH4 emissions from
agricultural practices including potential mechanisms behind observed eects. Furthermore, we investigate alterna-
tive uses of biochar in agricultural land management that may signicantly reduce the GHG-emissions-per-unit-of-
product footprint, such as (i) pyrolysis of manures as hygienic alternative to direct soil application, (ii) using biochar
Journal of Environmental Engineering and Landscape Management, 2017, 25(2): 114–139 115
Introduction: Human impact on global N2O and CH4
budgets and atmospheric concentrations
Human population now approaches 7.5 billion people
on earth. e land area that serves human nutrition and
bioenergy demands is not only limited, but declining due
to soil degradation in various forms (Lal 2014; Konuma
2016; FAO 2015). e dawning perception that “fertile
soils” are a nite global resource is stressed by recent land
grabbing practices where wealthier countries with large
populations and/or a lack of soil resources buy arable land
in poorer countries. is has mostly occurred in Africa
(Rulli etal. 2013), causing land use change to increase
crop productivity, likely with consequences in terms of
increasing GHG production. In fact, excessive human
land use change over the past decades has contributed to the
rapid, on-going increase in the atmospheric concentration of
non-CO2 greenhouse gases nitrous oxide (N2O) and methane
(CH4) from preindustrial levels of 270–280 ppbv to 324 ppbv
(N2O) and from ~700 ppbv to 1834 ppbv (CH4) (Myhre etal.
2013; Saunois etal. 2016).
e steep increase in atmospheric N2O concentra-
tions dominantly since the 1950s is clearly the result of
an increasing use of the Haber-Bosch process to gener-
ate reactive N forms from atmospheric N, plus the higher
proportion of legumes on farmland compared to natural
ecosystems; many important crop or fodder plants are N2-
xing legumes (e.g. soy, pea, lentils, beans, groundnuts,
clover). Global reactive-N use is now annually more than
double the amount introduced by natural processes (Gal-
loway etal. 2008); with increasing N fertilizer use comes
the increasing likelihood of N2O formation and atmo-
spheric accumulation. us, as stated by Ravishankara
etal. (2009), as the impact of uorinated halocarbons
decreases, N2O will likely become the dominant O3-de-
pleting substance within the agricultural sector over the
course of the 21st century.
Methane (CH4) emissions have also increased, by
150% since 1750 (Myhre etal. 2013) to 1834 ppb in 2015
(Dlugokencky etal. 1994; Saunois etal. 2016). Human
activities directly and indirectly contribute to the in-
creased atmospheric CH4 concentration by several path-
ways such as (i) expanding rice agriculture, ruminant
animal husbandry and landlling with unmanaged CH4
emissions, (ii) thawing permafrost areas and thermocast
lakes (Koven etal. 2011; Walter etal. 2006) and warming-
induced changes in plant community composition e.g.
expanding aerenchymal plant cover (Christensen etal.
2004), and (iii) “automatic” feedback eects such as rising
CH4 production under elevated, rising atmospheric CO2
concentrations from wetlands and agricultural lands due
to higher net biomass production as labile substrates for
methanogenesis (Van Groenigen etal. 2011).
To our knowledge, the last assessment of the human
impact of land-use changes and fertilizer use on global
CH4 consumption was made 20 years ago (Ojima etal.
1993). e authors estimated that human activities have
already reduced the global net CH4 sink capacity by 30%.
Hypotheses for explaining the reduction encompass hu-
man impact on (i) soil moisture changes, (ii) reduced soil
aeration via compaction / increased bulk density (both
impact gas diusivity and hence CH4 and O2 supply, (Cas-
tro etal. 1994; Hiltbrunner etal. 2012), and (iii) inhibition
by NH4+ (N fertilization or reduced nitrication in acidic
soils (Schnell, King 1995; Steudler etal. 1989). Also, (iv)
shis in the microbial community composition are hy-
pothesized, but without conclusive evidence (Gulledge
etal. 1997).
Over the last decades, the rising use of mineral N fer-
tilizer (Galloway etal. 2008), soil degradation and forest
clear-cutting, and a rising frequency of weather extremes
(Hansen, Sato 2016) creating “too wet” or “too dry” soil
conditions (Dijkstra etal. 2011) will likely further reduce
the global methanotrophic CH4 sink, and increase CH4
and N2O emissions. erefore, any positive contribution
that science and material use may provide agricultural
soils for reducing N2O emissions per unit yield, and in
reducing either the CH4 production and surface emis-
sions from soils or ruminant guts (see sections 3 and 4),
or by strengthening the soil CH4 oxidation capacity (sec-
tion 3.2) and its biolter function (section 3.1), needs to
be explored. Agricultural practices that utilize biochar
for meeting these needs appears promising, and should
be explored and developed to help lower the greenhouse
gas (GHG) footprint per unit yield or bioenergy produced
(also termed GHG intensity) (Wollenberg etal. 2016).
A reduction of N2O and CH4 within agroecosystems
can be achieved by either increasing the per-hectare yield
at unchanged GHG emission rates, by lowering the per-
hectare N2O and/or CH4 emissions/increasing CH4 up-
take, or ideally by both. Currently literature suggests that
biochar may play a role in reducing both of these GHGs;
biochar can aect GHG emissions directly following its
application to soils, and indirectly by adding carbonized
instead of non-carbonised residue or manures which usu-
ally have higher emissions following application. us, the
aim of this paper is to explore our current understanding
as fertilizer carrier matrix for underfoot fertilization, biochar use (iii) as composting additive or (iv) as feed additive in
animal husbandry or for manure treatment. We conclude that the largest future research needs lay in conducting life-
cycle GHG assessments when using biochar as an on-farm management tool for nutrient-rich biomass waste streams.
Keywords: biochar, greenhouse gases (GHG), nitrous oxide (N2O), methane (CH4), soil aeration, nitrate, soil N trans-
formations, GHG intensity.
C. Kammann et al. Biochar as a tool to reduce the agricultural greenhouse-gas burden – knowns, unknowns...
116
and knowledge gaps of biochar use as a tool to reduce N2O
and CH4 emissions from agricultural land use. To this
end, the biogeochemical mechanisms of formation and
consumption of N2O and CH4 in soils and their emission
to the atmosphere are presented and discussed. e topic
of build-up of the soil organic carbon stocks by biochar
amendment, or by reduced soil organic carbon decompo-
sition (negative priming) is not explored here, only some
rough assessments are made to illustrate potentials (e.g.
for using biochar as animal feed additive).
1. Eects of biochar application to soils on N2O
emissions
1.1. Mechanisms of N2O formation in soils and release
to the atmosphere
Soils are a prominent source of N2O emissions, especial-
ly when fertilised with organic or mineral N fertilisers.
A wide range of microbial and chemical processes and
pathways are responsible for these emissions (see Fig. 1),
with knowledge on these pathways continuously improv-
ing. e classical view was that bacterial denitrication,
the reduction of nitrate and nitrite in several steps to N2,
is the main source of N2O from most soils, especially at
intermediate water contents (Bateman, Baggs 2005). Vari-
ous bacteria that are phylogenetically unrelated are ca-
pable of denitrication, with many not having the full set
of enzymes for the complete pathway (Zum 1997). is
leads to the escape of intermediates, including N2O. Un-
der certain circumstances such as low pH and high NO3/
Corg ratios, the production of N2O is favoured compared to
the nal production of N2. e reduction of N2O to N2 is
performed by the enzyme nitrous oxide reductase, which
is encoded by the nosZ gene in denitrifying bacteria. In a
laboratory study, lower N2O emissions were inversely cor-
related to the nosZ gene expression (Harter etal. 2014).
Meanwhile, we know that bacterial nitriers may
dominate N2O emission from some soils, e.g. by the path-
way nitrier denitrication (Kool etal. 2011a). Bacterial
nitriers use ammonia as a substrate and reduce interme-
diately produced nitrite in a comparable way as in denitri-
cation. Nitriers have been found to be also able to use
nitrite provided exogenously in incubation studies under
aerobic conditions (Shaw etal. 2006). Fungi may play a
dominant role for N2O production from soils, either by
codenitrication or by fungal denitrication (Laughlin,
Stevens 2002; Rohe etal. 2014). e role of archaea for
N2O production from soils is still unknown, but there is
evidence of potentially signicant archaeal contributions
(Jung etal. 2014), although the responsible pathways are
yet unresolved (Jung etal. 2014; Stieglmeier etal. 2014).
As these processes may take place simultaneously in
dierent soil microsites, it is not straightforward to distin-
guish among them. Several methods have recently been
developed that try to unravel the sources of N2O, including
stable isotopes (Sutka etal. 2006; Kool etal. 2011b; Rohe
etal. 2014; Lewicka-Szczebak etal. 2016) and molecular
or modelling approaches (Rütting, Müller 2007; Kozlowski
etal. 2014; Perez-Garcia etal. 2014; Snider etal. 2015).
So far, no single method has oered a complete picture of
the diverse N2O producing pathways and a combination of
methods seems most promising.
1.2. Impact of biochar on soil N2O emissions: frequent
observations and assumed mechanisms
One of the rst biochar experiments reporting reduced
N2O emissions was presented in the 3rd USDA Symposium
on greenhouse gases and carbon sequestration in agricul-
ture and forestry (Rondon etal. 2005). A signicant de-
crease in N2O emissions was observed in pots planted with
soybean and grass in a greenhouse experiment. However,
Fig. 1. a) Processes and b) pathways of N conversions associated with N2O emissions in soils in relation to N transformations (A)
and N and C substrates (B). (Figure credit: N. Wrage-Mönnig)
a) b)
Journal of Environmental Engineering and Landscape Management, 2017, 25(2): 114–139 117
this nding went unnoticed for several years and only aer
the pioneering studies by Yanai etal. (2007), Spokas etal.
(2009) and van Zwieten etal. (2009), the number of pub-
lications on this topic started to rise. Hence, a new eld
of research was established, exploring a potential win-win
situation: biochar not only sequestered carbon but also
had the potential to decrease non-CO2 GHG emissions.
To date, the hypotheses for biochar’s impact in char-
soil mixtures on N2O emissions has been linked to bio-
char properties, the soil and the environmental conditions
such as temperature and precipitation (Spokas, Reicoscoky
2009; Dicke etal. 2015). Studies have mostly been carried
out in the lab using sieved/disturbed soil samples wetted
either to the same gravimetric moisture, or to the same
water-lled pore space, water holding capacity or water
potential. Other investigations in combination with plant
growth in the greenhouse or under less controlled con-
ditions in the eld have also shown that biochar may af-
fect the soil N2O emissions (e.g. Taghizadeh-Toosi etal.
2011; Schimmelpfennig etal. 2014; Kammann etal. 2012;
Deng etal. 2015; Hüppi etal. 2015). In the presence of
N2O-producing earthworms (soil fauna interactions), N2O
emissions were also reduced by biochar application (Au-
gustenborg etal. 2012; Bamminger etal. 2014). However,
in some studies no dierence between biochar and con-
trol treatments was observed (Scheer etal. 2011; Sánchez-
García etal. 2016) or N2O emissions were increased from
biochar amended soils (e.g. Spokas, Reicoscoky 2009;
Clough etal. 2010; Saarnio etal. 2013; Troy etal. 2013).
However, laboratory results cannot be generalised to eld
expectations. In eld trials, oen no statistical dierences
are observed between biochar and control treatments fol-
lowing eld application of biochar (Castaldi etal. 2011;
Jones et al. 2012; Karhu etal. 2011; Scheer et al. 2011;
Schimmelpfennig etal. 2014; Suddick, Six 2013; Dicke
etal. 2015). One potential reason for no signicant bio-
char eects on N2O emissions may be the application
dose, less homogeneous particle distribution and greater
soil (and plant) heterogeneity in elds resulting in high
variability in N2O uxes (large error bars, e.g. Hüppi etal.
2015).
Nevertheless, overall, meta-analyses conrmed that
N2O emissions are reduced with biochar application rates
of 1–2% by weight (van Zwieten etal. 2015; Cayuela etal.
2014). In spite of the extensive literature published during
the past several years on the topic, knowing if a biochar
will be eective in mitigating N2O emissions in a certain
agricultural eld is still highly unpredictable. us, most
research eorts are now directed towards achieving the
largest N2O emission reductions (what type of biochar to
use in what soils) by analysing the mechanisms involved.
Many studies have shown that biochar N2O mitigation
capacity will depend not only on the characteristics of
the biochar, but also on the type of soil and predominant
environmental conditions (Cayuela etal. 2013; Malghani
etal. 2013; Nelissen etal. 2014). A remarkable nding was
that, under identical environmental conditions, the same
biochar could increase emissions in one soil and decrease
emissions in another (Yoo, Kang 2012; Sánchez-García
etal. 2014). is fact seems to be linked to diverse N2O
formation mechanisms operating in dierent soils, of
which biochar might be aecting dierently. In this sense,
knowing how biochar interacts with the key microbial
pathways regulating N2O formation and consumption in
soil is crucial for developing and implementing eective
mitigation strategies. However, the number of studies
looking at specic N2O formation pathways is still very
limited and the mechanisms mediating N2O suppression
are still unresolved.
To date, most N2O-biochar studies selected certain
environmental conditions and assumed or speculated the
predominant N2O formation mechanisms. For instance,
studies at high water-lled pore space anticipated that the
main N2O formation pathway would be heterotrophic de-
nitrication. However, this reasoning has frequently been
shown to be incorrect. In complex soil environments,
ammonia oxidation and nitrier denitrication generally
coexist with heterotrophic denitrication (Hu etal. 2015)
and the proportion of N2O produced in each pathway de-
pends on many factors, not just water-lled pore space
(Wrage etal. 2001; Butterbach-Bahl etal. 2013). ere-
fore, studies that really dierentiate among N2O produced
by dierent sources aer addition of biochars are still ur-
gently needed.
1.2.1. What do we know about the impact of biochar on
denitrication N2O?
Denitrication is classically the most well-known mecha-
nism leading to N2O emissions and to date, also the most
investigated in biochar studies. Biochar might interact with
denitrication in dierent ways. Biochar might directly
stimulate or suppress total denitrication, i.e. the amount
of N that goes to gaseous form (N2 + N2O + NO). e
impact of biochar on total denitrication has been barely
studied and the results are inconclusive. For instance, us-
ing stable isotope enrichment, Cayuela etal. (2013) ana-
lyzed the ux of total N denitried (N2 + N2O) at the peak
of N2O emissions and found that biochar decreased the
total denitricatory N eux in 9 out of 15 soils, but sig-
nicantly increased the ux in two soils. Obia etal. (2015)
measured NO, N2O and N2 by high resolution gas kinet-
ics under strictly anaerobic conditions and calculated the
maximum induced denitrication rate, which was found
to increase with one type of biochar (cacao shell), but not
with another biochar (rice husk) in an acidic soil. By using
the acetylene inhibition technique, Ameloot etal. (2016)
found a general decrease in total denitried N (N2 + N2O)
with biochar in a neutral soil. Biochar might therefore
C. Kammann et al. Biochar as a tool to reduce the agricultural greenhouse-gas burden – knowns, unknowns...
118
decrease or increase total denitrication depending on the
type of soil. More studies are needed to understand the
mechanisms behind these observations.
A decrease in total denitrication was initially at-
tributed to improved soil aeration following biochar ad-
dition (Zhang etal. 2010), a hypothesis that has been
rebutted by Case etal. (2012) who demonstrated that
soil aeration played a minimal role in N2O mitigation.
Furthermore, several studies used adjusted water con-
tents to account for increased water holding capacities
that oen arise when biochar is mixed into the soil (e.g.
light-weight porous biochars in sandy soils), to render
simple aeration eects unlikely (e.g. Kammann etal.
2012); still these studies observed signicant N2O emis-
sion reductions. Another hypothesis suggested a gen-
eral decrease in soil microbial activity as a consequence
of toxic compounds present in biochar. For example,
phenolic compounds and PAHs have been observed to
contribute to the reduction in N2O release from agricul-
tural soils (Wang etal. 2013a). However, Alburquerque
etal. (2015) demonstrated that this hypothesis was un-
founded, since the presence of PAHs at typical biochar
concentrations did stimulate, rather than inhibit, N2O
emissions. Moreover, even if PAH containing biochars
would reduce N2O emissions, they will denitely never
be used in soils under any countries’ soil and fertilizer
regulations. In addition, numerous studies used clean
biochars with hardly detectable traces of even the most
abundant PAH (naphthalene) and these studies still
showed reduced N2O emissions (study compilations in
Cayuela etal. (2014) and van Zwieten etal. (2015)).
Several studies pointed out that microbial or
physical or plant immobilization of NO3 in soil fol-
lowing biochar addition could signicantly contribute
to the reduction of soil N2O emissions (compilation of
N2O-biochar studies in van Zwieten etal. 2015). is
hypothesis is reinforced by recent research showing
that biochar is able to capture considerable amounts of
nitrate, which is only partly detectable with standard
methods and largely protected against leaching (Kam-
mann etal. 2015; Haider etal. 2016). Nitrate capture
may physically separate nitrate from denitriers and
thus reduce nitrate availability.
On the other hand, biochar might interact with the
denitrication process by modifying the ratio of denitri-
cation products (N2O/N2). us, biochar may decrease
the N2O/N2 ratio (Cayuela etal. 2013; Harter etal. 2014;
Obia etal. 2015), but many questions arise from this
nding. For instance, Obia etal. (2015) related this phe-
nomenon to the alkalinizing eect of biochar in soil.
However, the decrease in the N2O/N2 ratio has also been
found in alkaline soils (Cayuela etal. 2013), where an
increase in pH did not occur aer biochar amendment.
Harter etal. (2016) found that biochar addition led to
the development of functional traits capable of N2O
reduction, containing typical and atypical nosZ genes.
Following a dierent line of research, several recent ar-
ticles highlight the importance of biochar redox prop-
erties, which may have a bigger impact on soil biogeo-
chemical processes than previously thought (Prévoteau
etal. 2016). In this line, Quin etal. (2015) measured
N2O reduction by injecting 15N-N2O in sterilized soil
columns and demonstrated that biochar took part in
abiotic redox reactions reducing N2O to dinitrogen (N2),
in addition to adsorption of N2O. Despite the current
knowledge about the impact of biochar on denitrica-
tion, additional studies are highly needed to explore the
detailed response mechanisms of denitriers to biochar
amendment.
1.2.2. What do we know about the impact of biochar on
N2O from nitrication and other processes?
It has been described that gross nitrication rates could
be increased aer biochar amendment because of high-
er substrate availability for nitrifying bacteria (Nelissen
etal. 2012), and several studies have analysed the impact
of biochar on gross and net nitrication (Prommer etal.
2014). ere is also a potential that biochar addition may
increase nitrication (and with it, N2O formation via ni-
trication pathways; Figure 1 in systems (such as needle-
rich raw humus soils) due to the sorption of phenolic
compounds; the latter can block or reduce nitrication. A
signicant increase in nitrication was seen in boreal for-
ests aer biochar addition (DeLuca etal. 2006; Ball etal.
2010) where sorption of phenols on biochar was respon-
sible for increased nitrication rates. However, only a cou-
ple of studies distinguished among N2O emissions from
nitrication pathways (via ammonia oxidation or nitrier
denitrication) and other sources by using isotopic signa-
tures of N2O, inhibition techniques or molecular methods.
In a laboratory incubation, Sánchez-García etal. (2014)
found that the addition of biochar increased N2O emis-
sions from a calcareous soil and concluded that the N2O
formation pathway operating in the soil was nitrication
(probably nitrier-denitrication). In another study, Wells
and Baggs (2014) showed that the biochar inuence came
primarily via ammonia oxidation, not N2O reduction or
production by denitriers, and increased N2O emissions
by 27%.
Dedicated studies of biochar eects on other soil
sources of N2O are largely missing. At low soil pH values,
it has been observed that fungi produced N2O instead of
N2 through codenitrication in presence of other nitro-
gen compounds, such as azide, salicylhydroxamic acid,
nitrite and ammonium (Liiri etal. 2002). Since biochar
can contain azide as well as other compounds, biochar
Journal of Environmental Engineering and Landscape Management, 2017, 25(2): 114–139 119
additions could theoretically enhance these fungal code-
nitrication processes. ere are limited studies examin-
ing the functionality of the N2O suppression through use
of selective microbial inhibitors, and data did support
the role of a particular microbial group in the N2O sup-
pression (bacteria or fungal, Lin etal. 2014). However, a
soil pH increase caused by biochar addition could also
reduce N2O production from fungal codenitrication,
thus the net outcome is unknown. Microbial community
composition will likely play a role: Using an identical
biochar in laboratory incubations across a series of 10
dierent soils, omazini etal. (2015) observed a trend
for the biochar suppression that could be correlated to
the total soil microbial biomass in the original soil. e
knowledge of eects of biochar additions on various
microbial sources of N2O are still little understood and
partly contradictory. Clearly, more research is needed
to be able to design biochars for the purpose of N2O
emission reduction not only in soils, but also when us-
ing biochar in the management of N-rich agricultural
(fertilizer) materials such as manures or composts (see
sections 3.1 and 3.3).
1.2.3. Long-term eects of biochar addition and in old
charcoal-rich soils: what do we know?
It is also still unclear how long N2O emission reductions
may persist following biochar addition to soil; or if old,
black-carbon rich soils that undergo a change in their
physico-chemical properties (such as Amazonian or Af-
rican Dark Earths) will have a lower or higher potential
for mitigating N2O emissions compared to soils without
biochar. While a lab study reported that aged biochar
particles increased N2O emissions (Spokas 2013), the op-
posite was observed in an experiment with >100 year-old
charcoal particles from a kiln site (Kömpf 2013). Hage-
mann etal. (2016) reported that biochar still signicantly
suppressed N2O emissions in the third season in the eld
compared to the corresponding control eld site without
biochar. More data are slowly emerging on old charcoal-
rich soils (e.g. from historic kiln sites, Borchard etal.
2014a; Hardy etal. 2016, 2017), and more Dark Earth
sites besides those in the Amazon basin have now been
identied (e.g. in Liberia and Ghana, Solomon etal. 2016).
However, to our knowledge no experiments on N2O emis-
sions and soil N transformations have yet been carried out
on these long-term analogues compared to their adjacent
native, non-black-carbon soils. For the overall question if
using biochar does oer long-term benets regarding N2O
emission suppression, exploring long-term eects is likely
of great importance, since reducing peak emissions in the
rst years will only be a small part over the long-term.
Particularly, these longer- and long-term eects are com-
pletely underexplored, and deserve much more research
attention in the near future.
2. Eects of biochar application on soil CH4 uxes
2.1. Mechanisms of CH4 uxes: production and
consumption in soils and net release to the atmosphere
e two biotic processes that determine the net methane
(CH4) exchange between soils/ecosystems and the at-
mosphere are methane production by strictly anaerobic
methanogenic Archaea (Methanogens) and methane con-
sumption by methanotrophic bacteria (Methanothrophs).
Methane production takes place in all anoxic environ-
ments where organic carbon is microbially degraded
(Conrad 2007a, 2007b; Whalen 2005), for example in
peatlands, lake sediments, ooded rice elds, in landlls,
in the guts of ruminant animals, termites or Scarabaeidae
larvae (Hackstein, Stumm 1994; Kammann etal. 2009).
Methanogens derive their energy from H2 and carbon di-
oxide (CO2) or acetate, formate, methanol or other pri-
mary and secondary alcohols and methylated compounds
(Brasseur, Chateld 1991; Conrad 1999). Methanogenesis
is thermodynamically the least ecient process i.e. other
reduction processes outcompete CH4 production, if the
concentration of alternative electron acceptors, such as
nitrate (NO3-), sulphate (SO42-), iron (Fe(III)) and manga-
nese (Mn(IV)), is high in relation to the input of organic
substrates (Lovley, Phillips 1987; Oremland 1988; Conrad
1989). Spatial and temporal variation in CH4 can be large
(e.g. Saarnio etal. 1997; Juutinen etal. 2003): Spatial vari-
ation in CH4 ux within dierent microsites of the same
ecosystem (e.g. Saarnio etal. 1997) and between dier-
ent ecosystem types (e.g. Saarnio etal. 2009). Water table
and temperature are dominant controls on CH4 eux in
bogs and swamps whereas the eect of aquatic vascular
plants (aerenchyma “ventilation”) was the most important
in fens or rice paddies (Turetsky etal. 2014). Besides the
aerenchyma transport, CH4 can also be transported with
the transpiration water stream in swamp tree species as
shown by Terazawa etal. (2007). Many 14CO2 experiments
have shown that recently xed C is rapidly delivered from
plants to methanogens (e.g. Megonigal etal. 1999) but the
amount of exuded carbon is many times lower than that
delivered via litter formation (Saarnio etal. 2004).
Methane consumption in soils is also ubiquitous in
all terrestrial environments (Hütsch 2001; Seiler etal.
1984), and human land-use changes such as deforesta-
tion, ploughing and N fertilization reduce the soil CH4
sink (Powlson etal. 1997). Net CH4 consumption is due
to the activity of methanotrophic α- and γ-proteobacteria.
Most methanotrophs use CH4 as the sole carbon source
and need oxygen to be active (Conrad 2007a). In upland
soils, methane oxidation is largely determined by the soil
diusivity for CH4 and O2 (Castro etal. 1994, 1995). Ac-
cording to their CH4-oxidation kinetics, categories of
“high-affinity” and “low-affinity” methanotrophs are
oen dened (Duneld 2007). e rst group occurs
C. Kammann et al. Biochar as a tool to reduce the agricultural greenhouse-gas burden – knowns, unknowns...
120
dominantly in upland soils and can consume atmospheric
and sub-atmospheric CH4 concentrations (<1800 ppb),
while low-anity groups are found in anoxic environ-
ments (e.g. rice paddy soils) in the aerobic centimeters or
millimeters of topsoil or in the oxygenated plant rhizo-
spheres; these methanotrophs need higher CH4 concentra-
tions. CH4 consumption can provide a “biolter function
for environments with high CH4 production (rice paddies,
landll cover soils etc.).
2.2. Biochar eects on CH4 production and release in
net CH4-source soils
e interactions between biochar application to soil and
CH4 uxes are not well understood, with disparate litera-
ture results (Jeery etal. 2016; Song etal. 2016). Biochar
application to paddy or ooded soils have been shown to
increase (Yu etal. 2013; Zhang etal. 2012), decrease (Feng
etal. 2012; Khan etal. 2013b; Lin etal. 2015; Qian etal.
2014), or have no signicant eect on CH4 emissions (Xie
etal. 2013). In anaerobic environments, the labile C pool
of biochar may theoretically function as methanogenic
substrate, promoting CH4 production (Zhang etal. 2010).
However, the labile C pool of root exudates and root lit-
ter is by far larger, thus labile biochar C may only play a
role (i) initially, (ii) when the biochar has been produced
at low temperature (i.e., greater labile C fraction), (iii)
in bare/fallow soils without root carbon supply, and (iv)
when the biochar amount added is great (>40 t/ha) (Saar-
nio 2016). As an example, Zhang etal. (2012) did not ob-
serve increased soil CO2 eux over two consecutive rice
cropping years with 10–40 t ha-1 biochar amendments,
but signicantly reduced N2O emissions at increased CH4
emissions; thus labile biochar-C is an unlikely explanation.
Biochar was reported to also promote methanotrophic
CH4 consumption at oxic/anoxic interfaces in anoxic envi-
ronments. is lowered the net CH4 emissions by the “bio-
lter” function of bacterial (low-anity) CH4 oxidation
before it escaped to the atmosphere. When methanotro-
phic organisms increasingly oxidise CH4 in the presence
of biochar at the oxic/anoxic root interface, they lower the
amount of CH4 that can enter into the plants’ aerenchyma
to escape (Feng etal. 2012).
In a greenhouse mesocosm study with sewage sludge
biochar (high application rates of 5% and 10%), rice yield
increased while the paddy soil turned from a net CH4
source to a net CH4 sink; this occurred in both rice-plant-
ed and bare paddy soil. ese results may have been due
to the addition of electron-accepting ash substances or ni-
trate with the sewage sludge biochar. In biochar-amended
landll cover soil, an increased CH4 oxidation activity was
responsible for decreasing CH4 eux from greater landll
depths (Sadasivam, Reddy 2015; Reddy etal. 2014). Here,
the physico-chemical properties including air conductivity
were considerably increased by biochar, i.e. the biochar ef-
fect might have been to improve the O2 supply to meth-
ane oxidizers. A recent meta-analysis (Song etal. 2016) of
CH4 emissions reported that biochar application caused
no pronounced change in CH4 emissions overall but
that there was signicant increase in methane emissions
(+19%). In another recent meta-study including papers
up to December 2015, Jeery etal. (2016) reported that
biochar amendment to ooded and/or acidic soils had the
potential to signicantly reduce CH4 emissions. ese two
meta-studies dier in their conclusions, which may be due
to a dierent database and meta-analytical approach.
Biochar impacts on natural net-methanogenic en-
vironments such as salt ats and wetlands are even less
well understood. Owing to the longevity of biochar and its
potential mobility and migration from anthropogenic sys-
tems, it will likely migrate to coastal areas over the long-
term (as recently shown for dissolved pyrogenic organic
carbon, Jaé etal. 2013). Lin etal. (2015) investigated bio-
char application to saline costal soils where soybean and
wheat was grown. ey did not nd any signicant eect
of biochar application to such soils on the (overall low)
CH4 emissions, but they observed a yield increase of 24
and 28% in soy and wheat, respectively. In rice eld stud-
ies, reductions in CH4 and/or N2O emissions were also of-
ten accompanied with increases in crop yields by between
10 and 20% (Dong etal. 2015; Khan etal. 2013a; Zhang
etal. 2012), resulting in a reduced greenhouse gas inten-
sity per kg of rice grain.
Biochar implementation may also reduce the GHG
intensity per unit of agricultural product by reducing
N-fertilizer and labile-C inputs at unaltered or increased
yields. Qian etal. (2014) reported that the use of 4 dif-
ferent biochar-compound fertilizers made of chemical
fertilizers, biochar and bentonite, at rates of well below
1t biochar ha–1, signicantly improved the GHG intensity
of a rice crop (by 36–56%) by: 1) the biochar-compound
fertilizer increasing grain yields by 10–31%; and 2) re-
ducing CH4 emissions by 25–50% and N2O emissions by
17–39%. ese results coincided with a reduced overall
N fertilizer input, from 210 kg N to 168 kg N ha–1, and
without taking the CO2-equivalents of reduced fertilizer
production and use into account (Qian etal. 2014). us,
there is a considerable potential for reducing the GHG
intensity of rice crop production, particularly in acidic
soils (Jeery etal. 2016), and that this potential extends
beyond the C-sequestration potential. Large amounts of
biochar do not need to be used at once since improve-
ments were found at biochar rates <1 t ha–1 (Qian etal.
2014). erefore, three central research topics emerge
here: (1) More research is needed on mechanisms of CH4
(and N2O) emission reductions with biochar use from
paddy and anaerobic soils including the eects on the
“methanotrophic biolter”; (2) Dedicated research on
Journal of Environmental Engineering and Landscape Management, 2017, 25(2): 114–139 121
biochar-eect-mechanism systematics to design biochars
with desired properties (omazini etal. 2015); and per-
haps most importantly (3) Research on biochar compound
fertilizers or underfoot fertilizers (Schmidt etal. 2015), to
achieve higher yields at reduced GHG emissions and re-
duced N fertilizer use.
2.3. Biochar eects on net CH4 consumption in oxic
upland soils
Well-aerated upland soils are characterised by CH4 con-
sumption, mediated by methanotrophic bacteria. Forest,
grassland and arable land have been described as CH4
sinks with ux rates of up to 65 µg CH4 m–2 hr–1 (Dalal
etal. 2008; Kern etal. 2012; Wang etal. 2005). As out-
lined above, in oxic/anoxic soil interfaces with a consid-
erable CH4 source strength, where the methane oxidiser
community is dominated by low-anity methanotrophs,
signicant increases in methanotrophic abundance and/or
activity have been reported following biochar amendment
(Feng etal. 2012; Reddy etal. 2014; Sadasivam, Reddy
2015). However, upland soils mostly host high-anity
methanotrophs, capable of consuming atmospheric meth-
ane. Species composition and biology is dierent to that
from anoxic soils (Duneld 2007), and the CH4-consum-
ing activity is quite easily hampered by human “activities”
such as land conversion (particularly deforestation), N
fertilization and ploughing. e number of studies using
upland soils and measuring CH4 consumption with/with-
out biochar is currently not large. Kollah etal. (2015) ob-
served signicantly increased CH4 consumption rates in a
lab study with tropical soil amended with biochar (with or
without organic amendments), as did Karhu etal. (2011)
in in-eld boreal, ploughed grassland soil. Karhu etal.
(2011) assumed that the observed doubling of the CH4
consumption in the ploughed grassland soil was due to
altered gas diusivity and water holding capacity (which
increased by 11%). However, the eects may also be con-
nected to increased N mineralization which usually occurs
aer ploughing, and where biochar may have prevented
N (NH4+) inhibition by sorbing NH4+ (Taghizadeh-Toosi
etal. 2012). Schimmelpfennig etal. (2014) observed in-
creased CH4 consumption in clay loam soil under labora-
tory condition; however, in the eld this was only present
in tendency. Similarly, Scheer etal. (2011) did not observe
increased soil CH4 consumption in a subtropical pasture
that had been amended by 10 t ha–1 manure biochar in
Australia.
In their meta-analysis, Song etal. (2016) reported
high levels of uncertainty for CH4 oxidation in upland
soils, while Jeery etal. (2016) concluded that biochar
addition may reduce the CH4 sink in neutral to alkaline
upland soils. Over all data sets, biochar had a CH4 sink-
increasing (or source-decreasing) eect in soils fertilized
at rates <120 kg N ha–1. Translated to upland soils this
indicates that, when true high-anity methanotrophic
activity is present, it may be increased by biochar applica-
tion. However, when high N application rates are used this
was not the case (Jeery etal. 2016); with high N fertiliza-
tion the CH4-oxidising activity of an agricultural soil is of-
ten considerably reduced or completely shut down, likely
due to nitriers replacing methanotrophs. In this case, the
methanotrophic population would not be supported or
improved. Taken together, the eects of biochar amend-
ment on soil CH4 consumption are not well understood.
Here, mechanistic studies are missing in particular, and
thus should be a focus for future research.
3. GHG emission reduction in animal husbandry and
waste management using biochar
Aer nearly a decade of research where biochar was
solely added to soil to assess GHG uxes, there is a shi-
ing perception, that biochar may also be used as a tool
to achieve GHG emission reductions during the handling
and management of organic nutrient-rich materials such
as manures. is section focuses on the use of biochar in
animal husbandry and in composting or plant-substrate
production (the topic of peat replacement is addressed
by Kern etal. 2017, this issue). In Germany, Austria and
Switzerland about 90% of the traded biochar is used in
animal husbandry, mainly as feed additive (in the way ac-
tivated carbon is used). However, to date, this topic has
been nearly neglected in biochar research.
3.1. Biochar as additive for feed and manure treatment
in animal farming to reduce the emission of GHG
Charcoal has been used to treat digestive disorder in
animals for several thousand years. Cato the Elder
(234–149 BC) mentioned it in his classic On Agriculture:
“If you have reason to fear sickness, give the oxen before
they get sick the following remedy: 3 grains of salt, 3 lau-
rel leaves, […], 3 pieces of charcoal, and 3 pints of wine.
(Cato 1935, §70; O’Toole etal. 2016). At the end of 19th to
beginning of the 20th century, charcoal was increasingly
used on a regular base to increase animal performance and
health (PSAC 1905; Day 1906; Savage 1917; Totusek, Bee-
son 1953; Volkmann 1935). Later during the last century,
veterinarian research focused on activated charcoal trials
mostly in the form of time-restricted medications against
intoxication and bacteriological as well as viral diseases
(Toth, Dou 2016; Schmidt etal. 2016). Only since about
2010 has biochar increasingly been used as regular feed
additive in animal farming (O’Toole etal. 2016), usually
mixed with standard feed at approximately 1% of the daily
feed intake. While scientists and farmers gained most of
the results and experience in cattle and chicken farming,
biochar is also administered to sheep, goats, pigs, horses,
rabbits, cats, dogs and extensively in sh farming (Toth,
C. Kammann et al. Biochar as a tool to reduce the agricultural greenhouse-gas burden – knowns, unknowns...
122
Dou 2016; Schmidt etal. 2016). In a German review pa-
per, Schmidt etal. (2016) evaluated more than 100 scien-
tic papers on feeding biochar to various animal groups.
Most of the studies showed for all investigated livestock
species, positive (but not always signicant) eects on pa-
rameters such as toxin adsorption, digestion, blood val-
ues, feed-use eciency, cell numbers in milk and livestock
weight gain; the latter may result from the pH-increasing
eect of various biochars since these are mostly alkaline
in nature (see 3.3). Buering the pH in the rumen could
likely prevent acidosis which is known to impact livestock
weight gain. However, only a small number of researchers
measured GHG emissions.
While chicken, pigs, sh and other omnivore ani-
mals provoke GHG emissions (mainly CH4, N2O; and
NH3) when their liquid and solid excretions decompose
anaerobically, ruminants cause direct methane emissions
through atulence and burps (eructation). is is espe-
cially the case for cattle that emit daily 500 to 600 l gas
with an average methane content between 6 and 8%. e
earliest evidence that feeding of biochar might reduce
cattle methane emissions came in a 2012 Vietnam study
(Leng etal. 2012a). In-vitro studies revealed signicant
methane reductions of 10 to 12.7% when biochar was fed
at rates between 0.5 and 1%. If biochar was blended with
nitrate, methane emissions were reduced by up to 49%.
Nitrate becomes a strong electron acceptor in the anaero-
bic rumen, keeping the hydrogen potential low, and thus
replacing a function of methane producing microorgan-
isms. Vongsamphanh etal. (2015) also found in in-vitro
tests with 1% biochar, while using cassava in rumen uid,
a 7% reduction of methane emissions within 24 h. In-vivo
trials of Leng etal. (2012b) revealed a reduction of enteric
methane of cattle by 20% with feeding 0.6% biochar and
by 40% with feeding 0.6% biochar blended to 6% potas-
sium nitrate, leading to a highly signicant animal weight
increase of 25% over 98 days. is is by far the most spec-
tacular result in reducing enteric cattle methane, but it has
unfortunately not yet been supported by other in-vivo or
in-vitro trials. Hansen etal. (2012) published the results of
an in-vitro trial with various non-characterized biochars
and their eect on methane production in rumen liq-
uids. All tested biochars showed a trend to reduce meth-
ane emissions between 11% and 17%. Other groups have
not repeated Leng and colleagues’ promising results. e
reason could be that Leng and colleagues used high tem-
perature gasier biochars made from rice husk, resulting
usually in carbonaceous materials with high electric con-
ductivity and electron buering capacity (Yu etal. 2015)
which may have had a stronger inuence on the digestion
reactions electrochemically than woody biochar. Using
biochar feeding to reduce ruminant methane emissions is
currently only an interesting perspective that needs more
systematic research.
Methane adsorption capacity by biochar is typically
the most investigated pathway for explaining eects when
fed to animals, but adsorption cannot explain all observa-
tions. Another decisive complementary function of bio-
char is its electro-biochemical interaction with biological
active systems, with research only recently beginning. Bio-
chars that are produced at temperatures above 700 °C are
not only good electrical conductors (Yu etal. 2015; Mo-
chidzuki etal. 2003) but can take part in biotic and abiotic
redox-reactions as an electron mediator (Husson 2013;
Kluepfel etal. 2014; Joseph etal. 2015a; Liu etal. 2012;
Shi etal. 2016; Van der Zee, Cervantes 2009; Yu etal.
2015; Kappler etal. 2014). A well balanced animal feed
regime contains multiple electron mediating substances,
however, in the high energetic diets of intensive livestock
farming these compounds are oen not contained in suf-
cient amounts (Sophal etal. 2013). If in these cases inert
or other non-toxic electron mediators like biochar, wood
vinegar or humic substances are added to the feed, many
redox reactions may take place more smoothly and e-
ciently which could increase energy conversion eciency
and thus feed eciency (Liu etal. 2012; Leng etal. 2013),
and eventually decrease enteric and post digestive GHG
production. Particularly, lowering CH4 emissions (which
are always a sign of energy loss) may be aided by the elec-
tron shuttling abilities of biochar. Moreover, it might be
assumed that the buering of the redox-potential as well
as the eect of electron shuttling between various micro-
bial species has a selective, milieu forming eect which
facilitates and accelerates the formation of functional mi-
crobial consortia and syntrophic species (Kalachniuk etal.
1994). e latter could explain why several studies found
a strong increase of Lactobacilli or a decrease of gram-
negative bacteria (Naka etal. 2001; Choi etal. 2009) which
seems to improve animal health. us, it may be hypoth-
esized that direct electron transfers between dierent spe-
cies of bacteria or microbial consortia (Chen etal. 2014)
via a biochar mediator may aid in a more energy ecient
digestion and thus higher feed eciency and eventually
result in lower GHG emissions (Leng etal. 2012a, 2012b).
When animals receive charcoal feed additives com-
bined with Lactobacilli spraying (i.e. microbial milieu
management in the stable), it is interesting to note that
antibiotic use may be reduced and in some cases down
to zero. Farmers in Germany who use this practice fre-
quently report reduced veterinarian costs (Kammann,
pers. comm.) that “pay” for the use of biochar and Lacto-
bacilli solution. Reduced antibiotics may also reduce CH4
emissions from ruminant husbandry. Recently, Hammer
etal. (2016) showed that application of broad-spectrum
antibiotics enhanced CH4 emissions from cattle manure,
and altered the gut microora from dung beetles feeding
on the manure from cows treated with broad-spectrum
antibiotics. As Choi etal. (2009) and Islam etal. (2014)
Journal of Environmental Engineering and Landscape Management, 2017, 25(2): 114–139 123
showed that feeding 0.3 to 1% biochar could replace anti-
biotic treatment in chicken and ducks, respectively, feed-
ing biochar plus administering Lactobacilli could have
an indirect eect on GHG emissions when it is able to
replace regular antibiotic “feeding”. Furthermore, Joseph
etal. (2015b) demonstrated that feeding biochar to graz-
ing cows had positive secondary eects on soil fertility
and fertilizer eciency, reducing mineral N-fertilizing re-
quirements which could be construed as another indirect
biochar GHG mitigation eect. us, enabling farmers to
stop or reduce administering antibiotics by using biochar
and Lactobacilli may be promising, not only for animal
health, but also for reducing methane emissions from ani-
mal husbandry operations (Hammer etal. 2016).
3.2. Calculating CO2-equivalent balances of biochar
use in animal husbandry: rst considerations
Besides the possible eects of biochar feeding on ruminant
CH4 emissions, it is not unlikely that microbial decom-
position of manure containing digested biochar produces
less ammonia, less CH4 and thus retains more nitrogen.
is has been observed between manure composted with
and without biochar (section 3.3; e.g., Sonoki etal. 2013;
Steiner etal. 2010; Troy etal. 2013; Wang etal. 2013b) and
may also occur when biochar is used as bedding or ma-
nure treatment additive. Ghezzehei etal. (2014) estimated
that using biochar for liquid manure treatment could save
57,000 t NH4 and 4,600 t P2O5 fertilizer per year in Califor-
nia alone, though this estimate is only based on laboratory
adsorption tests and not on eld trials. However, it cannot
be excluded that digested biochar will not have the same
eect on microbial decomposition, GHG emissions and
plant nutrient retention as when production-fresh biochar
is applied to the bedding or manure pit. To our knowl-
edge, there are no published data on GHG-emissions in
animal housing and of manure pits aer feeding animals
with biochar.
Easier to calculate is the C-sequestration potential of
biochar that is rst fed to livestock and eventually applied
to soil with the manure. Assuming an average C-content
of fed biochar of 80%, as required by the EBC feed certi-
cate (EBC 2012) and produced at recommended tempera-
tures above 500°C resulting in H/Corg ratios below 0.4, at
least 56% of the dry weight of the fed and manure-applied
biochar will persist as stable carbon in soil for at least
100years (Lehmann etal. 2015). If the global livestock
would, just theoretically and for the sake of a “back-of-the
envelope” assessment to explore magnitudes, receive 1%
of their feed in form of such a biochar, about 400 Mio.t
of CO2eq or 1.2% of the global CO2 emissions could be
compensated (Table 1).
While the feeding of “vegetal carbon” (biochar) is
permitted in the EU (EU 2011), it certainly cannot be
recommended yet as in a generalized biochar-livestock
feeding regime, since feed-grade certication of biochar
is currently not established in most countries and since
long-term eects are not suciently investigated. How-
ever, the potential for improving animal health and nu-
trient eciency, for reducing enteric methane emissions
as well as GHG emissions from manure management,
and for sequestering carbon while improving soil fertility
improvements, calls for increasing the scientic eort to
investigate, measure and optimize the GHG reduction po-
tential of biochar use in animal farming systems. e use
of biochar in animal husbandry is one of the largest unex-
plored research topics within the biochar research realm.
Although many unknowns and open questions exist, bio-
char use in animal operations appears promising from a
GHG reduction standpoint and thus future research could
focus eorts towards this area.
Table 1. Carbon sequestration potential of biochar fed to livestock (globally) with subsequent manure application to soil. Numbers
of total livestock follow FAO statistics obtained for 2014 (FAO 2016). An intake dosage of 1% of the daily feed weight was assumed.
e C sequestration potential was calculated with the assumption of 80% C in biochar, and 70% C persistence over 100 years
(Lehmann etal. 2015; Camps-Arbestain etal. 2015)
Animal Global number of
individual animals Daily intake dosage*
(g bc animal–1 d–1)Annual intake dosage
(kg bc animal–1 yr–1)Total biochar
(Mio. t yr–1)Total C seq.
(Mio. t yr–1)Total C O2eq
(Mio. t yr–1)
Cattle 1,482,144,415 120 43.8 64.9 36.4 133.3
Bualoes 195,098,316 120 43.8 8.5 4.8 17.5
Sheep 1,209,908,142 50 18.25 22.1 12.4 45.3
Goats 1,006,785,725 50 18.25 18.4 10.3 37.7
Pigs 986,648,755 80 29.2 28.8 16.1 59.2
Horses 58,913,957 120 43.8 2.6 1.4 5.3
Chickens 21,321,834,000 6 2.19 46.7 26.1 95.9
Turkeys 461,453,000 6 2.19 1.0 0.6 2.1
Total 193.0 108.1 396.3
C. Kammann et al. Biochar as a tool to reduce the agricultural greenhouse-gas burden – knowns, unknowns...
124
3.3. Biochar as composting additive
Composting is the aerobic biotic oxidation of organic res-
idue, manure and waste for producing organic fertilizer
and occasionally waste management. Labile organic car-
bon is transformed into humic like substances and CO2,
with N2O and CH4 as non-desired by-products. Emission
rates of N2O and CH4 depend on compost management,
and increase with increasing degrees of anoxia during
composting. During a perfect aerobic quality compost-
ing procedure, N2O and CH4 emissions are usually low
(Amlinger etal. 2003, 2008; Bernal etal. 2009). However,
aerobic conditions cannot always be perfectly maintained,
particularly when manure-rich waste is composted; here
larger N2O and CH4 emissions are common (Li etal. 2016;
Wang etal. 2013b). us, it is desirable to develop strate-
gies to reduce GHG emissions during composting, par-
ticularly of nutrient-rich wet materials.
With regard to biochar, “co-composting” refers to the
addition of biochar to the initial composting feedstock.
During co-composting, biochar can sorb compost liquids
rich in nutrients, particularly nitrate (Prost etal. 2013;
Kammann etal. 2015). Furthermore, some studies report
an accelerated thermophilic phase with higher compost
temperatures (Kammann etal. 2016; Vandecasteele etal.
2016). CO2 emissions are an obligate result of compost-
ing due to the decomposition of labile organic matter. is
loss during composting ranges from ~40% (Vandecasteele
etal. 2016) to ~80% (Sánchez-García etal. 2015) depend-
ing on the compost feedstock and composting conditions.
Some studies show no eect of biochar on CO2 emissions
or increases of emission rates (Steiner etal. 2011; López-
Cano etal. 2016). However, Malinska etal. (2014) and
Vandecasteele etal. (2016) reported a reduction of CO2
emissions. Vandecasteele etal. (2016) argue that pristine
biochar used in their experiments might have physically
adsorbed CO2 (Creamer etal. 2014; Fornes etal. 2015)
and thus reduced CO2 emissions, although actual CO2
production probably increased (e.g., increase of decom-
position, higher temperatures). However, a faster rate of
C loss does not necessarily mean that, overall, biochar will
reduce the long-term storage of non-pyrolyzed feedstock
in soils; it just means that a process that happens other-
wise over longer periods is accelerated.
CH4 and N2O emissions from compost are non-de-
sired side eects and can be reduced by adequate man-
agement of the composting process with regard to oxygen
supply, including windrows not exceeding certain sizes,
optimized water content and forced or mechanical aera-
tion (Fukumoto etal. 2003; Amlinger etal. 2008). Howev-
er, these management strategies can be costly. Recent stud-
ies indicate that biochar addition during the composting
process can indeed reduce emissions of both CH4 and N2O
probably due to enhanced access of oxygen mediated by
biochar. Methane emissions are oen drastically reduced
by biochar addition, with reported reduction rates of 55%
for chicken manure compost with biochar added at 20%
w/w (Jia etal. 2016), >70% for organic waste compost with
biochar added at 10% rate (Sonoki etal. 2011), and >80%
for municipal solid waste compost with biochar added at
10% w/w (Vandecasteele etal. 2016). However, at low bio-
char additions (e.g. 3%), Sánchez-García etal. (2015) ob-
served no signicant eect on CH4 emissions during the
composting process, suggesting that a certain biochar rate
during composting is necessary to reach desired reduc-
tions in CH4 emissions.
Gaseous N losses also oen decrease when biochar is
used as an additive during the composting process since
the pH of composts does usually stay below or around
Fig. 2. Mean GHG ux rates of compost (comp), biochar-compost (BC-comp) and compost with later addition of fresh biochar
(comp + BC), n = 5 + standard deviation; GHG ux measurements with 400 g substrate per 1-L Weck® jar as described in
Kammann etal. (2012); Material properties and composting procedure described in Schmidt etal. (2014) and Kammann etal.
(2015). Substrates were adjusted to 60% of their respective water-holding capacity one week before measurement. Letters indicate
signicant dierences by one-way ANOVA (n = 5, p < 0.05)
CO2 efux (mg CO2 kg–1
h–1)
N2O emission (ng N2O kg–1
h–1)
CH4 ux (mg CH4 kg–1
h–1)
Journal of Environmental Engineering and Landscape Management, 2017, 25(2): 114–139 125
pH 7.5. However, care has to be taken to prevent a strong
pH increase due to (large amounts of) biochar addition
in composting, with animal bedding, manure treatment
or elsewhere. If values above a pH of 8 are reached, NH3
losses may result which was systematically investigated
by Chen etal. (2013) with biochar additions up to 20%
by weight to bauxite processing sands in a pH-adjusted
set-up of pH values ranging from 5 to 9. e authors re-
ported that at low and high pH values, NH3 losses were
increased by biochar additions (at low pH when the ad-
dition of biochar was suciently large to provoke a pH
increase), while at a high pH NH3 losses were high any-
way. However at a more neutral pH, the NH4+ adsorption
capacity of biochar dominated which decreased N losses
(Chen etal. 2013). However, the rst available results are
contradictory: While Schimmelpfennig etal. (2014) found
a reduction of NH3 emissions when slurry was added to a
loam soil (pH 6.0) amended with biochar when compared
to the same soil amended with (less alkaline) straw, Subedi
etal. (2015) reported an increase in the NH3 emissions
from slurry when both, an alkaline biochar and an acidic
hydrochar were added to the slurry. It is possible that the
overall “evaporation surface” which may increase by add-
ing a bulky material to soil or slurry may also play a role
for increasing NH3 losses. Nevertheless, the biochar-dose-
to-pH-increase relationship should always be taken into
account.
Studies conducted with relatively low dosages of bio-
char (e.g. 3–4%) showed either a reduction in N2O emission
(Wang etal. 2013a; Li etal. 2016) or no eect (Sánchez-Gar-
cía etal. 2015; López-Cano etal. 2016). When reductions in
N2O emissions were observed, they were pronounced but
happened only during a portion of the composting process.
For example, Li etal. (2016) reported that 3% biochar ad-
dition during composting reduced N2O emission by 54%,
which was entirely attributable to N2O-peak suppression for
only one of the eight measurement dates. is eect was
further attributed to a marked reduction in the abundance
of the nirK gene of denitrifying bacteria when biochar was
co-composted (Li etal. 2016).
Ammonia (NH3) is not a greenhouse gas, but a rele-
vant atmospheric precursor of N2O (shown in section1.2),
and its volatilization during composting is a relevant path-
way for N loss. Biochar amendment was shown to reduce
NH3 volatilization particularly in N-rich materials such as
sludge or manures, probably due to NH4+ sorption (Chen
etal. 2010; Hua etal. 2009; Steiner etal. 2010; Malins-
ka etal. 2014). Increased N content in the compost, in
the form of NH4+, is a desired property for compost use
as an organic fertilizer. When applied to soil, the larger
quantity of N and labile C retained in the composted bio-
char particles (or in the biochar-compost product) may
theoretically lead to higher N2O emissions as compared to
pure, non-composted biochar (Prost etal. 2013; Borchard
etal. 2014b). is was, however, not observed in a plant-
ing study using co-composted (nitrate-enriched) biochar
as soil amendment (Kammann etal. 2015), or when the
compost itself was tested (Fig. 2).
At the moment, results are just snapshots of rst
composting experiments with biochar, and studies in-
clude a wide range of experiments from 150 g mixtures in
jars (Jia etal. 2016), 45 L composting reactors (Malinska
etal. 2014), 1 m³ gardening compost boxes (Prost etal.
2013) to 60×3×2 m windrows with forced aeration and
automated mechanical turning (Vandecasteele etal. 2016)
or daily machine turning (Schmidt etal. 2014; Kammann
et al. 2015) at varying time scales (few weeks to sev-
eral months). Generally, the value of composting studies
would be increased if the resulting compost quality is sub-
sequently evaluated by plant germination and growth tests
(e.g. Hua etal. 2012, soil faunal tests, e.g. Fig. 3) and eld
application trials.
Biochar is a promising tool to optimize composting,
improve compost quality and charge biochar with plant-
available nutrients and reduce non-CO2 GHG emissions
during composting. However, the longer-term eects of
biochar-composts in soils are largely unexplored. More
systematic studies with dierent combinations of N-rich
and N-poor feedstock and dierent well-characterized
biochars under comparable, praxis-relevant conditions are
needed, and subsequently product quality and its eects
on soil GHG emissions aer soil amendment should be
investigated.
Fig. 3. Results of earthworm avoidance tests (method: ISO
17512-1:2008; Busch etal. 2012) with the compost (comp),
biochar-compost (bc-comp) and compost with later addition of
fresh biochar (comp+bc), material properties and composting
described in Schmidt etal. (2014) and Kammann etal. (2015).
“Comp vs. comp” is the control and did not result in an eect
(as needed); boric acid (BA) was the “positive control” where
avoidance behaviour conrmed that the test was valid. Bars
give means of n = 5 repetitions per treatment (50 worms),
shown as preference or avoidance of the respective substrate
(Fishers exact test, (*) p < 0.1, * p < 0.05, ** p < 0.01, n.s. not
signicant)
C. Kammann et al. Biochar as a tool to reduce the agricultural greenhouse-gas burden – knowns, unknowns...
126
4. Novel approaches of using biochar for GHG
emission reductions in agriculture
GHG emission reductions with biochar use may not nec-
essarily be direct; rather, reductions may result from al-
ternative waste stream management involving pyrolysis.
Another option is to use biochar as a fertilizer carrier to
match plant demands more closely, enabling the reduced
use of common mineral fertilizers and hence associated
N2O emissions (see 3.1, Qian etal. 2014). is section ad-
dresses novel biochar concepts for reducing the overall
GHG intensity of agricultural production, including those
from unwanted NO3- and NH3 losses and deposition, caus-
ing N2O emissions elsewhere (Turner etal. 2015).
4.1. Biochar as carrier for (organic) underfoot
fertilizers
e use of large biochar amounts (e.g. 10–20 tons) being
ploughed into soils does usually not increase crop yields
suciently to justify the costs (Jeery etal. 2015; Bieder-
man, Harpole 2013; Ruysschaert etal. 2016). Here, the
economy of biochar use may be better for special crops
as long as yield increases without negative eects on crop
quality can be realized, as observed by Baronti etal. (2014)
and Genesio etal. (2015) for grapevine cultures. However,
grapevines are low-N cultures with low N2O emissions
(Marras etal. 2015) compared to higher N-fertilizer de-
manding vegetable crops such as radish, carrot, spinach
or potato. ese receive much higher N fertilizer amounts,
oen with several cropping cycles per year, resulting in
high N2O emissions (Ruser etal. 1998; Min etal. 2016). In
highly fertilized vegetable crops, biochar amendment may
play a role in signicantly reducing the greenhouse gas in-
tensity per unit of crop, likely by reducing N2O emissions
and partly by yield increases (Jia etal. 2012; Li etal. 2015).
However, use of biochar in these situations, as with others,
comes with a cost.
In order to reduce costs, it is desirable to maximize
the eect of biochar per unit applied, suggesting that op-
timum application rates might be fairly low especially if
the product can be concentrated around the root zone.
Indeed, root zone application of only 2–4 t ha–1, together
with fertilizer in conservation farming systems, has been
reported to improve yields in acidic sandy soils (Cornelis-
sen etal. 2013). In addition, using biochar as an ingredi-
ent of, or carrier matrix for, mineral or organic fertilizer
blends (“fertichars”) has recently been investigated as a
promising strategy to reduce the need for large biochar
amounts while improving crop yields and/or reduce N
use (Joseph etal. 2013). Supporting this contention, Qian
etal. (2014) showed that biochar organo-mineral fertil-
izer blends reduced the GHG-intensity of rice crop pro-
duction even at reduced N fertilization rates and at very
low amounts of biochar (<1 t ha–1).
In practice, the on-farm availability of clean, inex-
pensive, and amounts required will decide whether bio-
char will be used, particularly in rural areas in developing
countries (Cornelissen etal. 2013). In Nepal, a self-made
organic “fertichar” was used as underfoot fertilizer in 8
dierent farmer sites (fertile loamy soils) and compared
to the same nutrient additions without biochar. Here,
pumpkin yields were consistently increased by 400% with
the urine-loaded biochar applied as underfoot fertilizer,
and multiple eld trials on dierent soils, crops and bio-
char consistently showed a growth-promoting eect of
the organic biochar-root zone fertilizer (Schmidt etal.
2015). It has not yet been tested if these biochar-fertil-
izer blends will reduce GHG emissions compared to an
equal or greater amount of standard fertilizer. However,
we argue that GHG emission reductions (namely those
of N2O) per unit of crop yield may be achieved for the
following reasons: (1) e fertichar blends mean a com-
paratively high amount of biochar concentrated together
with the nutrients, i.e. within the “biochar concentration
range” found to reduce N2O emissions in soils consider-
ably in meta-studies (Cayuela etal. 2014); (2) By stimu-
lating plant growth, N uptake increases, which reduces
the availability of N for nitrication / denitrication and
hence N2O formation; (3) When applied underfoot, the
total land area that receives N fertilizer and is able to pro-
duce N2O emissions is smaller; (4) By using root-zone fer-
tichars, the overall amount of N fertilizer that needs to be
applied, per unit of crop produced, can likely be reduced
(Joseph etal. 2013; Qian etal. 2014). is may automati-
cally reduce the GHG emissions (as a percentage of the
applied fertilizer); and (5) Biochar will reduce the soil bulk
density (e.g. Obia etal. 2016) and increase soil air capac-
ity, which has been shown to reduce N2O production due
to improved oxygen supply in upland soils. However, it is
simply unknown if this type of application will reduce the
GHG intensity of crop production and if this will occur on
dierent soils and climatic zones alike. Further research
is needed on fertilizer-biochars, either as broadcast gran-
ules (Qian etal. 2014) or as underfoot fertilizers (Schmidt
etal. 2015), since these techniques may oer a win-win
in terms of reducing GHG emissions concomitantly with
improved yields.
4.2. Pyrolysis as alternative waste stream management
technique
Over the past twenty years, localization of conned ani-
mal feeding operations (CAFOs) in the USA has brought
about a massive production of animal manures (Gollehon
etal. 2001). Traditional management practices have ap-
plied manure to crops as a fertilizer, yet long-term manure
applications have caused nutrient imbalances in soils far
in excess of what crops can assimilate (Barker, Zublena
Journal of Environmental Engineering and Landscape Management, 2017, 25(2): 114–139 127
1995; Kellogg etal. 2000). e situation is so severe that
CAFO producers must have nutrient management plans
(e.g., MDE 2016; ISDA 2016) or manure application can
be restricted if soils contain excess N and/or P. A clear
need exists for alternate animal waste management and
recycling methods.
ermal processing of animal manures using gasi-
cation and pyrolysis technology is gaining considerable
interest as an alternate treatment option because of the
energy quantity generated (Cantrell etal. 2007, 2008).
Manure gasication is a popular conversion process, yet
the higher temperatures (>900 °C) required for conver-
sion into gases, along with impurities in the feedstocks
(e.g. salts, silicates, etc.), oentimes limits the conversion
eciency (Lv etal. 2010) and can corrode downstream
metal surfaces (Demirbas 2005). us, pyrolysis is more
oen regarded as the eective method of processing bio-
mass to produce a combination of non-condensable gases,
bio-oil, and biochar (Antal, Grønli 2003). Manures can be
pyrolyzed at various temperatures (e.g. 300 to 750°C) to
create biochars. Aerwards, the energy content (as MJ/kg)
contained within biochar identies the quantity of energy
generated per equivalent dry weight and serves as a con-
venient energy index relative to coal (Table 2). Globally,
animal waste pyrolysis is performed to create a thermal
energy source, to generate heat for animal connement
stables, for the production of bio-oil, and to produce bio-
char, a nutrient-enriched end-product to be used as a soil
amendment (Laird etal. 2009; Lee etal. 2013) or fertilizer
replacement.
Table 2. Net thermal energy in various manures, and their
biochars pyrolyzed at various temperatures, as compared to
hard/so coals, gasoline and methane. na = not available
Feedstock ermal
tempe ra-
ture (°C)
Energy
content
(MJ/kg) Source
Poultry litter 0 15 Novak etal. (2013)
700 14.2 Novak etal. (2013)
Dairy manure
0 17.6 Cantrell etal.
(2012)
350–700 19.0–20.9 Cantrell etal.
(2012)
Swine Solids
0 19.4 Cantrell etal.
(2012)
350–700 15.1–21.1 Cantrell etal.
(2012)
Human Feces 300–750 13.8–25.6 Ward etal. (2014)
Hard/so coal na 29.3–33.5 Euronuclear (2016)
Gasoline na 43–47 Energynumbers
(2005)
Methane na 55.5 Energynumbers
(2005)
Following manure pyrolysis, it has been shown by
Gaskin etal. (2008) that nutrient availability may be de-
creased in the biochar, which may make manure-based
biochar use attractive in terms of land areas where nu-
trient management plans are necessary; biochars may
be able to supply a more balanced quantity of essential
plant nutrients without degradation in environmental
quality due to nutrient over-application. Furthermore,
an additional benet similar to previously mentioned
underfoot fertilizer-biochars, manure-based biochar
may significantly reduce overall farm-management
based GHG emissions simply by reducing mineral fer-
tilizer use (Cayuela etal. 2014; Nguyen etal. 2014).
e production of animal manure biochar via py-
rolysis, and the subsequent partial or complete replace-
ment of standard mineral NPK fertilization by manure-
biochar, oers the following pathways to N2O emissions
reduction: (1) Reduced N use which reduces the overall
environmental burden by limiting N2O formation both
from direct manure application as well as emissions
originating from o-site N pollution (e.g. NH3 emis-
sions, N export via overland ow, N leaching to ground-
water, Turner etal. (2015); (2) e use of manure-based
biochar may reduce N2O emissions compared to using
the same amount of N in the mineral or manure form.
Cayuela etal. (2014) examined 107 articles related to
manure-based biochar land application, showing a –46
to +39% change in N2O emissions, with average N2O
emission changes close to 0%. us, changes or reduc-
tions in N2O emissions is likely process dependent,
something that was not examined by Cayuela et al.
(2014). However, Subedi etal. (2016) added poultry lit-
ter or swine manure biochar to two dierent soils. e
authors showed that poultry litter or swine manure bio-
chars pyrolyzed at 400 °C produced the same N2O emis-
sion factor (i.e., N2O emitted as a percentage of the total
N supplied) as control soils. When pyrolyzed at 600 °C,
the N2O emission factor decreased for both biochars as
compared to controls. Higher temperatures during pro-
duction will result in manure-based biochars containing
lesser quantities of easily degradable C compounds (e.g.,
volatile compounds) that are available for denitriers,
leading to less likelihood of anaerobic soil conditions
(e.g., Liu etal. 2014) and thus lower N2O emissions. Ob-
viously, in order to reduce N2O emissions when apply-
ing manure-based biochars, easily degradable C sources
needs to be at a minimum; and (3) e use of pyrolysis
for heat generation in animal housing, and/or the use of
manure biochar for energy production, can be imple-
mented to reduce GHG emissions when it replaces the
use of fossil fuels for the same purpose (see energy con-
tent comparison in Table2).
C. Kammann et al. Biochar as a tool to reduce the agricultural greenhouse-gas burden – knowns, unknowns...
128
Conclusions: promising options for GHG emission
reduction and future research needs
After a decade of intense biochar research, it has be-
come clear that biochar soil amendments are able to
reduce N2O emissions (i.e. emission peaks). Biochar
can also reduce CH4 emissions, particularly in flooded
soils, and when N fertilization rates are not too high.
However, great uncertainty still exists with respect to
biochar use and its GHG reducing effect as associated
with different biochars and soil types/conditions. This
is due to the lack of understanding of mechanistic bio-
char effects. Good “candidate” mechanisms that might
explain N2O emission reductions are pH increases (lim-
ing effect) and changes in microbial community com-
position, particularly changes in the denitrifier gene
expression and abundance, and N (mostly nitrate) cap-
ture in biochar particles. It is likely that these mecha-
nisms work in concert under field conditions. Good
candidate” mechanisms that might explain CH4 emis-
sion reductions are the stimulation of methanotrophic
low-affinity communities at the anoxic/oxic interface in
reduced environments as well as the electron shuttling
and redox activity of biochars (e.g. rice paddies and ru-
minant guts).
Biochar use in animal husbandry is economically
promising and the dominant route of biochar use in
central Europe; however, research on GHG emission
reductions is largely lacking. Studies dealing with the
medical use of charcoal/biochar in animal feeding near-
ly never characterize char properties. Biochar shows a
strong potential for reducing GHG and NH3 emissions
in the composting of wet nutrient-rich materials, par-
ticularly when composts are not so frequently turned
(aerated, i.e. “lazy composting”). Yet, more detailed
studies on the dynamics of this effect along the com-
posting process are needed, as well as a cost-benefit
evaluation for potential implementation.
One new promising option of using biochar for
improving the GHG-intensity (yield-to-GHG emission
ratio), via both increasing yields plus reduced GHG
emissions, may be the use of designed (organic) “Fer-
tiChars” (biochar as fertilizer carrier) administered as
concentrated root zone fertilizers. Here the biochar-
to-N ratio may certainly be in the range where biochar
should have an effect as well as effectively and envi-
ronmentally deliver optimal nutrient concentrations to
plants. However, to date, GHG flux measurements of
this implementation pathway are lacking and strongly
call for future research. We conclude that the use of
biochar in agriculture provides a unique opportunity to
reduce the non-CO2 greenhouse gas “cost” per unit of
yield produced, yet future research is required to maxi-
mize its benefits.
Acknowledgements
We thank the participants of the ANS & EU-COST Bio-
char Symposium 2015 in Geisenheim for their active and
lively discussions in section 6 “Biochar and GHG emis-
sions”. C. Kammann and H.-P. Schmidt thank Matthias
Schröder for his work during the composting trials in Ar-
baz, Switzerland and later in the lab at Gießen University.
Contribution
CK developed the structure, wrote sections, contributed to
sections and coordinated writing. JI and NH helped stream-
lining and nalizing the manuscript in addition to section
contributions. All other authors contributed equally by re-
viewing the literature within their particular eld of exper-
tise and writing sections of the manuscript. (erefore, all
authors besides authors 1–3 are listed in alphabetical order.)
Funding
e COST Action TD1107, “Biochar for sustainable en-
vironmental management”, provided nancial support for
the EU-COST biochar symposium in 2015; the German
BLE and FACCE-JPI funded the German participants
of the “DesignerChar4Food” (D4F) project (CK: Project
No.2814ERA01A; NW-M: Project No. 2814ERA02A),
the Spanish colleagues (JME and TFM) were funded by
FACCE-CSA nº 276610/MIT04-DESIGN-UPVASC and
IT-932-16, and US colleagues (JN, JI and KS) were funded
by e USDA-National Institute of Food and Agricul-
ture (Project # 2014-35615-21971), USDA-ARS CHAR-
net and GRACENet programs – D4F greatly stimulated
discussions. We also gratefully acknowledge the follow-
ing individual grants that enabled authors to contribute
to this work: Claudia Kammann: German Science Foun-
dation (DFG) grant [KA3442/1-1]; ML Cayuela thanks
Fundación Séneca (Grant number 19281/PI/14); Nils
Borchard was placed as integrated expert into the Center
for International Forestry Research (CIFOR) by the Cen-
tre for International Migration and Development (CIM).
CIM is a joint operation of the Deutsche Gesellscha für
Internationale Zusammenarbeit (GIZ) GmbH and the In-
ternational Placement Services (ZAV) of the German Fed-
eral Employment Agency (BA).
Disclosure statement
None of the authors do have any competing nancial, pro-
fessional, or personal interests from other parties.
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... When used in animal farming systems, biochar also reduces environmentally harmful ammonia loss through volatilization or nitrate loss via leaching (Emauel & Ernest, 2021;Borchard et al., 2019;Sha et al., 2019) and can reduce greenhouse gas emissions, such as nitrous oxide (N 2 O) (Kammann et al., 2017;Hegarty et al., 2024;Borchard et al., 2019) or methane (CH 4 ) (Jeffery et al., 2016). ...
... Biochar was mixed with animal feed for the first time in Germany and Switzerland in 2012 (Graves et al., 2022). A substantial portion of industrially produced biochar in Europe since then has been dedicated to sales for applications in animal feed, bedding, manure treatment, and subsequent soil use (Kammann et al., 2017;Rathnayake et al., 2023). In 2016, acknowledging the surge in biochar use within animal-related sectors, the European Biochar Foundation introduced a distinct certification standard tailored for animal feed (EBC, 2018). ...
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... Studies also report reduced N 2 O emissions from biochar-amended soils [166,172]. While biochar has been shown to decrease CH 4 emissions [173], its ability to reduce greenhouse gases is variable due to factors such as biochar type and soil conditions [173]. Biochar's nitrate sorption capacity depends on factors like pyrolysis temperature and surface area, with some biochars showing higher sorption rates [174,175]. ...
... Studies also report reduced N 2 O emissions from biochar-amended soils [166,172]. While biochar has been shown to decrease CH 4 emissions [173], its ability to reduce greenhouse gases is variable due to factors such as biochar type and soil conditions [173]. Biochar's nitrate sorption capacity depends on factors like pyrolysis temperature and surface area, with some biochars showing higher sorption rates [174,175]. ...
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... Additionally, N2O emissions are probable to rise by 35-60% in the next few years This increase is mostly attributable to inefficient organic waste handling and amplified use of fertilizers made up of chemicals (Haider et al., 2020). In addition, improper timing as well as excessive nitrogen use can cause N leaching, reducing water quality and increasing N2O emissions from landscape-draining rivers (Kammann et al., 2017). N2O is predominantly generated in soil by microorganisms converting reactive Nitrogen. ...
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... For quite some time, char has been used as a feed supplement for domesticated animals, and the outcomes of this supplement for various animal species have generally been positive. This corresponds to a better metabolism, a higher feed conversion ratio, weight gain, a decrease in greenhouse gas emissions, a cure for intoxication, and treatment for bacterial or viral infections [94]. ...
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