Chemosphere 252 (2020) 126539

Contents lists available at ScienceDirect

Chemosphere
journal homepage: www.elsevier.com/locate/chemosphere

Review

Biochar technology in wastewater treatment: A critical review

Wei Xiang a, b, Xueyang Zhang a, b, *, Jianjun Chen c, Weixin Zou d, Feng He e, Xin Hu f,
Daniel C.W. Tsang g, Yong Sik Ok h, Bin Gao b, **
a
School of Environmental Engineering, Jiangsu Key Laboratory of Industrial Pollution Control and Resource Reuse, Xuzhou University of Technology,
Xuzhou, 221018, China
b
Department of Agricultural and Biological Engineering, University of Florida, Gainesville, FL, 32611, USA
c
Mid-Florida Research & Education Center, University of Florida, Apopka, FL, 32703, USA
d
Jiangsu Key Laboratory of Vehicle Emissions Control, Nanjing, 210093, China
e
College of Environment, Zhejiang University of Technology, Hangzhou, 310014, China
f
Center of Material Analysis, Nanjing University, Nanjing, 210093, China
g
Department of Civil and Environmental Engineering, The Hong Kong Polytechnic University, Hong Kong, China
h
Korea Biochar Research Centre & Division of Environmental Science and Ecological Engineering, Korea University, Seoul, South Korea

h i g h l i g h t s g r a p h i c a l a b s t r a c t

Biochar technologies in various

wastewater treatment are elucidated.

Feedstock pre-treatment and post-
treatment effect on biochar produc-
tion is reviewed.

Biochar as an innovative adsorbent to
remove aqueous contaminants is
discussed.

Future perspectives of biochar tech-
nology in wastewater treatment are
summarized.

a r t i c l e i n f o a b s t r a c t

Article history: Biochar is a promising agent for wastewater treatment, soil remediation, and gas storage and separation.
Received 27 January 2020 This review summarizes recent research development on biochar production and applications with a
Received in revised form focus on the application of biochar technology in wastewater treatment. Different technologies for
11 March 2020
biochar production, with an emphasis on pre-treatment of feedstock and post treatment, are succinctly
Accepted 17 March 2020
summarized. Biochar has been extensively used as an adsorbent to remove toxic metals, organic pol-
Available online 18 March 2020
lutants, and nutrients from wastewater. Compared to pristine biochar, engineered/designer biochar
Handling Editor: X. Cao generally has larger surface area, stronger adsorption capacity, or more abundant surface functional
groups (SFG), which represents a new type of carbon material with great application prospects in various
Keywords: wastewater treatments. As the first of its kind, this critical review emphasizes the promising prospects of
Engineered biochar biochar technology in the treatment of various wastewater including industrial wastewater (dye, battery
Wastewater treatment manufacture, and dairy wastewater), municipal wastewater, agricultural wastewater, and stormwater.
Production technologies Future research on engineered/designer biochar production and its field-scale application is discussed.
Modification methods
Based on the review, it can be concluded that biochar technology represents a new, cost effective, and
Carbonaceous adsorbents
environmentally-friendly solution for the treatment of wastewater.
© 2020 Elsevier Ltd. All rights reserved.

* Corresponding author. School of Environmental Engineering, Jiangsu Key Laboratory of Industrial Pollution Control and Resource Reuse, Xuzhou University of Technology,
Xuzhou, 221018, China.
** Corresponding author.
E-mail addresses: zhaxuy@163.com (X. Zhang), bg55@ufl.edu (B. Gao).

https://doi.org/10.1016/j.chemosphere.2020.126539
0045-6535/© 2020 Elsevier Ltd. All rights reserved.
2 W. Xiang et al. / Chemosphere 252 (2020) 126539

Contents

1. Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2
2. Production technologies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2
2.1. Pre-treatment technologies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2
2.2. Thermal carbonization technologies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3
2.3. Post-treatment technologies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5
3. Biochar as an adsorbent for aqueous contaminant removal . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5
3.1. Heavy metal removal . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5
3.2. Organic contaminant removal . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7
3.3. Nitrogen and phosphorus removal . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7
4. Biochar technology in wastewater treatment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8
4.1. Industrial wastewater treatment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8
4.2. Municipal wastewater treatment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8
4.3. Agricultural wastewater treatment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9
4.4. Stormwater treatment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10
5. Conclusions and future perspectives . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11
Declaration of competing interest . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12
Acknowledgements . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12
References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12

1. Introduction shown in Fig. 1, pre-treatments vary depending on feedstock and

the purposes for biochar use, including physical (dry, crush, sieve,
Biochar is a porous carbonaceous material produced during the wash, etc.), chemical (treat with chemicals or functional materials,
thermochemical decomposition of biomass feedstock in the pres- load of precursors and functional agents, etc.), and biological
ence of little or no oxygen. Biomass feedstock can be any organic (bacterial treatment, etc.) methods. Post-treatments mainly rely on
waste materials which include crop and forest residues, wood chip, physical (ball milling, magnetization, etc.) and chemical (corrosive
algae, sewage sludge, manures, and organic municipal solid wastes treatment, etc.) methods (Zhang and Gao, 2013; Tan et al., 2016;
(Colantoni et al., 2016; Xiong et al., 2019). Methods for thermo- Usman et al., 2016). Thus far, only few review articles have
chemical decomposition include pyrolysis, hydrothermal carbon- emphasized pre-treatments in relation to feedstock decomposition
ization, gasification, torrefaction, and microwave heating, varying methods and resultant biochar properties as well as post-treatment
in thermochemical temperature and duration (Mohan et al., 2014; technologies on biochar properties and their effects on wastewater
Gonz alez et al., 2017; Fang et al., 2018). Interest in biochar is largely treatment.
based on its two distinct benefits: First, biochar production per se The overarching objective of this work is to present the first
can offset greenhouse gas emission because it stores carbon in a comprehensive review on the applications of biochar technology in
stable form, preventing the release of greenhouse gases into the wastewater treatment. After summarizing new technologies on
atmosphere from biomass degradation (Creamer and Gao, 2016; pre-treatment of feedstock, thermal carbonization process, and
Yang et al., 2018a). Second, biochar is an effective, low-cost, and post-treatment of biochar (Section 2), this review digests current
environment-friendly adsorbent (Cha et al., 2016; Inyang et al., knowledge of biochar as an innovative adsorbent for aqueous
2016), which is related to its relatively large surface area and contaminants (Section 3). Most importantly, recent advances of
abundant surface functional groups (SFG) (Wang et al., 2017a; biochar applications in wastewater treatments, including industrial
Zhang et al., 2017a). Biochar can be used for adsorbing metals/ wastewater, municipal wastewater, agricultural wastewater and
metalloids and purifying water (Agrafioti et al., 2013; Van Vinh stormwater are perspicuously and detailly elucidated (Section 4).
et al., 2015; Palansooriya et al., 2019), applied to soils for This critical review also discusses the perspectives and future
improving soil fertility and crop productivity (Yoo et al., 2018), research directions of the biochar technology in wastewater
employed for clean energy production to partially replace the fossil treatment (Section 5).
fuels (Fang et al., 2018; Cao et al., 2019), and utilized as adsorbent
and catalysts to various pollutants and reduce greenhouse gas
emission (Xiong et al., 2017). As a result, biochar becomes 2. Production technologies
increasingly important as a solution to some global problems, such
as climate change, environmental pollution, and soil degradation 2.1. Pre-treatment technologies
(Creamer and Gao, 2016).
It has been well documented that feedstock, thermochemical Pre-treatment is the first step for biochar production from
decomposition methods and their temperature and duration can different raw materials. In general, these methods can be classified
significantly affect biochar physical and chemical properties (Yu physical, chemical, and biological pre-treatment technologies
et al., 2019). Several previous review articles have discussed (Fig. 1).
decomposition methods, characterization, and applications of Physical pre-treatment technology generally includes drying,
biochar in removal of different contaminants from aqueous solu- crushing, sieving, and washing of biomass feedstock. The feedstock
tions (Mohan et al., 2014; Cha et al., 2016; Tan et al., 2016). Biochar riches in lignocellulosic/plant is usually dried to constant weight at
properties can also be affected by feedstock pre-treatments and 105  C or other temperature, ground into smaller particles using a
biochar post-treatments (Wang et al., 2017a; Yang et al., 2019). As hammer mill, and then cut into different pieces (Wang et al., 2016a;
Essandoh et al., 2017; Zhang et al., 2017a). Occasionally, separate
 W. Xiang et al. / Chemosphere 252 (2020) 126539 3

Fig. 1. Biochar production technologies: (a) Biomass. (b) Pre-treatment technologies. (c) Thermal processes. (d) Post-treatment technologies.

drying step may be needed for some plant feedstock, because the production of engineered biochar with enlarged surface area,
plant raw materials vary greatly in moisture contents. Physical pre- unique pore structure, enriched SFG, etc. (Zhou et al., 2017a; Zhao
treatment method for biomass feedstock is related to its own et al., 2018).
properties. For example, dewatered sludge is often dried in an oven Biological pre-treatment technology is a relatively new concept
overnight, crushed, sieved, and stored in sealed containers prior to that utilizing biological processes to improve biomass feedstock for
use (Agrafioti et al., 2013). Newspapers and cardboard are engineered biochar production (Wang et al., 2017a). Bacterial
commonly shredded and blended into pulp as the feedstock treatment, particularly anaerobic digestion or biofuel processes, of
(Randolph et al., 2017). Paper mill sludge is acid-washed, rinsed biomass feedstock has been proven to be an effective and product
with deionized distilled water to obtain mineral-free sludge (Cho ‘biologically activated’ biochar with enhanced properties (Inyang
et al., 2017). Algae is alkaline, it is usually rinsed with fresh water et al., 2010; Yao et al., 2015). In the literature, several biomass
and then dried, granulated or flaked before pyrolysis (Roberts and materials including sugar beet tailings, bagasse, sludge, and animal
de Nys, 2016). waste were subjected to the anaerobic digestion process first and
Chemical pre-treatment technology often relies on chemical then the residues were converted into biochar through slow py-
reactions to change the properties or compositions of feedstock rolysis (Inyang et al., 2010; Yao et al., 2011a; Tang et al., 2019). The
materials. One type of most commonly used chemical pre- anaerobic digestion pre-treatment would make the obtained bio-
treatment technique is to treat feedstock biomass materials with char have a larger specific surface area (SSA) and better adsorption
chemicals or functional materials to load chemical precursors or performance (Inyang et al., 2010; Yao et al., 2011a). It is recognized
functional agents into the feedstock. During the treatment, the that utilizing the biological pre-treatment residue materials to
biomass feedstock is immersed into a chemical solution or a produce biochar can introduce additional benefits such as reducing
colloidal suspension, and then dried prior to biochar production waste disposal costs, and making bioenergy more eco-friendly
(Tan et al., 2016). After pretreated with metal ion solutions such as (Inyang et al., 2010; Yao et al., 2015). Another biological pre-
FeCl3, AlCl3, and MgCl2, biomass feedstock can be successfully treatment method uses biomass enriched with high concentra-
converted into biochar-based nanocomposites with metal oxy- tions of minerals including heavy metals through bioaccumulation
hydroxide (e.g. Fe3O4, AlOOH, and MgO) nanoparticles stabilized on for biochar production (Yao et al., 2013b; Wang et al., 2017c). Wang
carbon surface with the pores of the engineered biochar (Zhang et al. (2017c) converted a heavy metal hyperaccumulating plant
et al., 2012a, 2013; Zhang and Gao, 2013; Son et al., 2018). On the into biochar and suggested that this technology not only provides a
other hand, biomass can be pretreated with engineered nano- safe solution for hyperaccumulator disposal but also produces
particles and natural colloids including carbon nanotubes, gra- value-added biochar nanocomposites.
phene and clay, which also leads to the successful production of
biochar-based nanocomposites (Zhang et al., 2012b; Yao et al., 2.2. Thermal carbonization technologies
2014; Inyang et al., 2015). Corrosive chemicals including acid, al-
kali, and oxidant have also been applied to pretreat biomass for the Thermal processes for biomass conversion into biochar mainly
4 W. Xiang et al. / Chemosphere 252 (2020) 126539

include pyrolysis, microwave-assisted pyrolysis, hydrothermal

carbonization and gasification (Mohan et al., 2014; Wang et al.,
2017a; Fang et al., 2018). Table 1 summaries and compares these
carbonization technologies.
Pyrolysis is a thermochemical process for decomposing biomass
in an anoxic or hypoxic environment (Cha et al., 2016). Pyrolysis
processes depend on the operating temperature, heating rate, and
residence time used, which can affect the compositions and phys-
icochemical properties of products. The yields of biochar decrease
with increasing pyrolysis temperature (Fig. 2), whereas ash and
carbon content increase (Fig. 3). It is mainly related to cellulose,
hemicellulose and lignin decomposition, saline-alkali separation,
carbonization and other factors in biomass (Hossain et al., 2011; Al-
Wabel et al., 2013; Masek et al., 2013; Irfan et al., 2016). The heating
rate determines the pyrolysis speed, and it influences the charac-
teristics of biochar and the yield of bio-oil and bio-gas (Inyang et al.,
2010; Cho et al., 2017). Prolonged residence time provides more
complete biomass decomposition while decrease the biochar pro-
duction yield (Mohamed et al., 2016).
Microwave-assisted pyrolysis (MAP) is considered as a sustain-
Fig. 2. Percent yields of biochar from different feedstock at different pyrolysis tem-
able method to produce bio-energy products, including biochar, perature (data are from reference (Yuan et al., 2011; Bian et al., 2016; Colantoni et al.,
bio-oil, and bio-gas (Dai et al., 2017; Mutsengerere et al., 2019). In lez et al., 2017)).
2016; Irfan et al., 2016; Lin et al., 2016; Wang et al., 2016a; Gonza
comparison to the conventional methods, MAP technique offers
shorter processing time, lower energy requirement, more effective
heat transfer, and better selective heating (Zhang et al., 2017b;
Duran-Jime nez et al., 2018). The MAP process is mainly controlled
by the microwave power, irradiation time, etc. (Lam et al., 2017;
Duran-Jime nez et al., 2018; Nhuchhen et al., 2018; Kadlimatti et al.,
2019). The yield of biochar often decreases as the microwave power
increases, which can be attributed to the high heating rates at high
microwave power levels (Jimenez et al., 2017; Nhuchhen et al.,
2018). Biochar with high SSA was obtained in a microwave sys-
tem operated at the microwave power of 500 W, irradiation time of
3 min, and frequency of 2450 ± 25 MHz. (Dura n-Jimenez et al.,
2018). Further microwave treatment, however, resulted in a loss
of SSA, which can be attributed to the degradation of micropore
structure of the biochar after the microwave overheating (Jimenez
et al., 2017).
Hydrothermal carbonization (HTC) is the conversion of wet
feedstock at a temperature range of 120e260  C into biochar
without pre-drying (Mohan et al., 2014; Cha et al., 2016; Fang et al.,
2018). The wet biomass is heated and pressurized (2e10 MPa) for Fig. 3. Percent of carbon and ash in biochar from different feedstock at different py-
5e240 min in a confined system (Kambo and Dutta, 2015; Fang rolysis temperature (data are from reference (Hossain et al., 2011; Al-Wabel et al.,
et al., 2018; Zhang et al., 2019a). The biochar produced by HTC is 2013; Masek et al., 2013)).

Table 1
Summary of common thermal carbonization technologies (Cha et al., 2016; You et al., 2017; Mutsengerere et al., 2019; Zhang et al., 2019b).

Thermal carbonization Key parameters Temperature/ Residence Desired Advantages

technologies power range time product

Pyrolysis temperature; 300e850  C 1e3 h Biochar Simple, robust, and cost-effective; applicable to small scale and farm-based
heating rate; biochar production
residence time
Microwave-assisted microwave power; 400e500 W 1e10 min Biochar and volumetric, fast, selective, and efficient heating
pyrolysis microwave biofuel
irradiation time
Hydrothermal temperature; 120e260  C 1e16 h Hydrochar More suitable for feedstock with high moisture content
carbonization residence time;
pressure;
water-to-biomass
ratio
Gasification temperature; ＞800  C 10e20 s Syngas Biochar yield of gasification is less than pyrolysis, but the biochar contains a
particle size; high level of alkali salts (Ca, K, Si, Mg, etc.).
residence time;
pressure;
gasification agent/
biomass ratio
 W. Xiang et al. / Chemosphere 252 (2020) 126539 5

usually called hydrochar. Reaction temperature is identified as the doped biochar has been successfully synthesized by simply ball
governing parameter during the HTC (Kambo and Dutta, 2015). milling pristine biochar with ammonium hydroxide, these N groups
With the increase of temperature, hydrochar contains abundant improve the adsorption performances of the biochar on acidic
acidic functional groups on its surface, which can benefit the carbon dioxide and anionic reactive red (Xu et al., 2019). Ball-
contaminant adsorption capability (Zhou et al., 2017a; Saha et al., milling technology is thus an effective engineering method to
2019). Increasing holding temperature and holding time can in- produce novel engineered biochar. The ball-milled biochar shows
crease the porous structure of the hydrochar, which increases the enhanced physicochemical and adsorptive properties, and can be
possibility of the application of hydrochar as an adsorbent (Shao used in various environmental applications.
et al., 2019). Corrosive treatments such as acid, alkali, and oxidation treat-
Gasification is the process converting the biomass to gas fuel ments are commonly used chemical modification techniques,
using gasification agents. Gasification temperature is generally which alter the surface chemistry of the biochar. The corrosive
higher than 800  C (You et al., 2017). The biochar produced during chemicals, such as HCl, HNO3, KOH, NaOH, KMnO4, and H2O2 have
gasification usually contains high levels of alkali salts and alkaline been applied to modify biochar for different purposes (Wang et al.,
earth mineral (Kambo and Dutta, 2015; Zhang et al., 2019b), which 2015a, 2017a; Cha et al., 2016; Zheng et al., 2019). The chemical
can precipitate many heavy metal contaminants and thus be used modified biochar has higher SSA, more microporous, more func-
directly as a remediation agent in problem soils (Yang et al., 2018b; tional groups, and enhanced sorption capacity (Yang et al., 2019).
Yu et al., 2019). Deal et al. (2012) reported that problem soils Alkali (NaOH)-acid (HNO3) combined modification shows an
amended with gasifier-produced biochar had higher maize yields, obvious increased BET surface area, porosity and oxygen-
and the soluble ash content of the biochar had the greatest influ- containing functional groups of municipal sewage sludge biochar,
ence on soil productivity. which enhances tetracycline adsorption, up to 286.9 mg/g (Tang
et al., 2018). KMnO4 and KOH treatment increase the SSA of bio-
2.3. Post-treatment technologies char derived from waste peanut shell, resulting in increased
adsorption sites for Ni2þ (An et al., 2019). H2O2 is another strong
Biochar are often post-treated by either physical or chemical oxidant for modifying biochar (Xue et al., 2012). H2O2-modified
modification methods to increase its SSA, pore volume, surface manure biochar can eliminate heavy metals efficiently, due to the
chemistry, and functional agents including SFG and composited increased oxygen and carboxyl group content (Wang and Liu, 2018).
nanoparticles (Van Vinh et al., 2015; Tan et al., 2016; Dai et al., Post-treatment of biochars represent a new area of research. It
2017). In the literature, there are several good reviews that have modifies existing biochars by increasing biochars’ SSA, pore vol-
provided comprehensive summaries of various post-treatment ume, negative zeta potential, oxygen-containing functional groups,
technologies for biochar modifications (Tan et al., 2016; Wang and the adsorption capacity. Such modified biochars can be cost-
et al., 2017a). This review thus only slightly discusses three post- effective and environmentally-friendly carbon materials with
treatment technologies including magnetic, ball milling, and cor- great application potential in many fields.
rosive (i.e., acid, alkali, or oxidation) treatment (Mohamed et al.,
2016; Usman et al., 2016; Wang et al., 2017a), which either are 3. Biochar as an adsorbent for aqueous contaminant removal
current research hotspots or have not reviewed intensively in the
literature. Biochar can be used as an adsorbent to remove different pol-
Magnetization is the method that converts biochar into a lutants in water and wastewater. Here, we mainly discuss its use for
magnetic material where magnetic iron oxides including Fe3O4, g- removal of heavy metals, organic contaminants, nitrogen and
Fe2O3, or CoFe2O4 particles are loaded into biochar (Zhang et al., phosphorus.
2013; Wang et al., 2015b; Tan et al., 2016; Shengsen Wang et al.,
2019). Thus, magnetic modified biochar can easily be recovered 3.1. Heavy metal removal
from the aqueous solution (Zhang et al., 2013; Mohan et al., 2014;
Wang et al., 2015b; Son et al., 2018). Magnetic zero-valent iron Heavy metals in wastewater can adversely affect human beings,
biochar derived from peanut hull at 800  C has a higher removal animals, and plants. Long term exposure to heavy metals in the
rate for Cr6þ, which is mainly due to its high SSA, pore volume, and aqueous phase can cause serious health threats even at low con-
loaded reductive iron (Liu et al., 2019b). Another method for pre- centration (Ahmed et al., 2016). Increased evidence suggests that
paring magnetic biochar composites is directly chemical co- biochar obtained from plants and animal residues can effectively
precipitate Fe3þ/Fe2þ on biochar surface (Tan et al., 2016). Mag- adsorb heavy metals in water and wastewater (Higashikawa et al.,
netic switchgrass biochar prepared by the precipitation of iron 2016; Inyang et al., 2016; Tan et al., 2016; Dai et al., 2017; Zhou
oxide using an aqueous Fe3þ/Fe2þ solution has the highest et al., 2017a). Table 2 summarizes biochar adsorption of heavy
adsorption capacity for metribuzin (205 mg/g, pH ¼ 2) (Essandoh metals in aqueous phase.
et al., 2017). Arsenic is an extremely toxic metal and occurs in wastewater as
Ball milling is a simple and efficient method which uses the well as drinking water. The adsorption capacity of As3þ is enhanced
kinetic energy by moving balls to break chemical bonding, chang- from 5.7 mg/g to 7.0 mg/g through the surface modification of bio-
ing the particle shape and producing nanoscale particles (Lyu et al., char by Zn(NO3)2 impregnation (Van Vinh et al., 2015). Biochar
2017). After ball milling, the characteristics of biochar were produced from paper mill sludge was applied to adsorb As5þ and
enhanced including SSA, pore volume, negative zeta potential, the maximum adsorptive capacity was 34.1 mg/g (Cho et al., 2017).
oxygen-containing functional groups, and the adsorption capacity Biochars produced separately from sugarcane straw, rice husk,
(Wang et al., 2017a; Lyu et al., 2018a, 2018b; Xiang et al., 2020). Ball- sawdust, and chicken manure were mixed with sawdust and used
milled bagasse biochar has higher Ni2þ removal efficiency than to remove Cd2þ in water. Results show that increased pyrolysis
pristine biochar, and the adsorption capacity of Ni2þ and aqueous temperature from 350  C to 650  C triggers the increasing tread in
methylene both increased (Lyu et al., 2018b). This is mainly due to percentage removal of Cd2þ (Higashikawa et al., 2016). Biochars are
the fact that ball milling can increase the external and internal also effective in removal of Pb2þ. The removal efficiencies of Pb2þ by
surface areas of the biochar and expose its graphitic structure and biochars produced from fresh and dehydrated banana peels are
oxygen-containing functional groups (Lyu et al., 2018a). Nitrogen- 359 mg/g and 193 mg/g, respectively (Zhou et al., 2017a). Table 2
6 W. Xiang et al. / Chemosphere 252 (2020) 126539

Table 2
Biochar adsorption of heavy metals in aqueous solutions.

Biochar Pre-treatment Thermal Post treatment Pyrolysis Biochar Adsorption Heavy Initial Adsorption Removal mechanism Ref.
feedstock process temperature dose (g/ pH metals concentration capacity
( C) L) (mg/L) (mg/g)

Bamboo Oven dried Pyrolysis HNO3þ nZVI 600 2 e Agþ 200 584 Innersphere Wang et al.
wood treated complexation and (2017b)
electrostatic attraction
by outer-layer Fe oxides
under oxic conditions
Bamboo Oven dried Pyrolysis H2O2þ nZVI 600 2 e Agþ 200 1217 Innersphere Wang et al.
wood treated complexation and (2017b)
electrostatic attraction
by outer-layer Fe oxides
under oxic conditions
Pomelo Dried þ H3PO4 Pyrolysis Pristine 250 2 6 Agþ 50 137.4 Chemical adsorption Zhao et al.,
peel impregnated with oxygenic (2018)
functional groups
Pine wood Oven dried and Pyrolysis Ni/Fe-LDH 600 2.5 7.5 As3þ 20 4.38 Electrostatic attraction Wang et al.
milled modified and surface (2016b)
complexation with
hydroxyl groups
Pine wood Ni/Fe-LDH Pyrolysis Pristine 600 2.5 7.5 As3þ 20 1.56 Electrostatic attraction Wang et al.
modified and surface (2016b)
complexation with
hydroxyl groups
Paper mill Oven dried and Pyrolysis Pristine 720 1 2.7e10.4 As5þ 26.7 34.1 Chemisorption or Cho et al.,
sludge acid washed chemical reaction (2017)
process between
available adsorption
sites and adsorbate
Sewage Stirred and Pyrolysis Pristine 300 4 e As5þ 0.05 e Chemical sorption Agrafioti
sludge heated et al.,
(2013)
Sewage Stirred and Pyrolysis Pristine 300 4 e Cr3þ 0.2 e Chemical sorption Agrafioti
sludge heated et al.,
(2013)
Rice husk Washed Pyrolysis Polyethylenimine 450e500 1 e Cr6þ 100 435.7 Introduction of amino Rajapaksha
modified group facilitate chemical et al.,
reduction of Cr6þ and (2016)
increase sorption
capacity
Green Dried Pyrolysis HCl modified 600 2 3e8 Cd2þ 5.6 6.72 Chemisorption Zhang et al.,
waste (2018)
2þ
Peanut shell Washed, dried Pyrolysis Hydrated 400 0.2 6.5 Cd 10 10 Nonspecific outer- Wan et al.,
and milled manganese oxide sphere surface (2018)
treated complexation provided
by oxygen-containing
groups, specific
innersphere
complexation offered by
the impregnated HMO
Marine FeCl3 Pyrolysis Pristine 500 16.7 e Cu2þ e 69.37 Oxygen-containing Son et al.,
macro- immersed functional groups as (2018)
algal potential adsorption
sites
Banana Oven dried Pyrolysis Pristine 600 2.5 e Cu2þ 200 75.99 Electrostatic attraction, Ahmad
peels partial of physisorption, et al.,
ion exchange and (2018)
precipitation
Cauliflower Oven dried Pyrolysis Pristine 600 2.5 e Cu2þ 150 53.96 Electrostatic attraction, Ahmad
leaves partial of physisorption, et al.,
ion exchange and (2018)
precipitation
Pomelo Dried þ H3PO4 Pyrolysis Pristine 250 2 6 Pb2þ 50 88.7 Precipitated by Zhao et al.,
peel impregnated phosphorous functional (2018)
groups
Peanut shell Washed, dried Pyrolysis Hydrated 400 0.2 6.5 Pb2þ 20 36 Nonspecific outer- Wan et al.,
and milled manganese oxide sphere surface (2018)
treated complexation provided
by oxygen-containing
groups, specific
innersphere
complexation offered by
the impregnated HMO
Banana Oven dried Pyrolysis Pristine 600 2.5 e Pb2þ 600 247.1 Electrostatic attraction,
peels partial of physisorption,
 W. Xiang et al. / Chemosphere 252 (2020) 126539 7

Table 2 (continued )

Biochar Pre-treatment Thermal Post treatment Pyrolysis Biochar Adsorption Heavy Initial Adsorption Removal mechanism Ref.
feedstock process temperature dose (g/ pH metals concentration capacity
( C) L) (mg/L) (mg/g)

ion exchange and Ahmad

precipitation et al.,
(2018)
Cauliflower Oven dried Pyrolysis Pristine 600 2.5 e Pb2þ 200 177.8 Electrostatic attraction, Ahmad
leaves partial of physisorption, et al.,
ion exchange and (2018)
precipitation
Maple Dried Pyrolysis H2O2 modified 500 5 7 Pb2þ 50 43.3 Complexation by Wang et al.,
wood oxygen functional (2018)
groups
Pecan Dried and MAP Pristine e 2 3 Pb2þ 500 80.3 Ion-exchange by Jimenez
nutshell milled calcium ions on the et al.,
material surface (2017)
Banana Dehydrated HTC Pristine 230 0.25 7 Pb2þ 200 359 Ions exchange and Zhou et al.
peels and grinded surface complexation. (2017a)
Banana H3PO4 soaked HTC Pristine 230 0.25 7 Pb2þ 200 193 Ions exchange and Zhou et al.
peels surface complexation. (2017a)
2þ
Peanut hull Dried HTC Pristine 300 2 e Pb 50 0.88 Complexation with Xue et al.,
carboxyl surface (2012)
functional groups
Peanut hull Dried HTC H2O2 modified 300 2 e Pb2þ 50 22.82 Complexation with Xue et al.,
carboxyl surface (2012)
functional groups

also shows biochar adsorption of Cr3þ, Ni2þ and Cu2þ. Biochar the effectiveness is closely related to the aromaticity index, polarity
prepared from sewage sludge adsorbed approximately 70% of Cr3þ index, SSA, and the quantity of oxygen functional groups (Mohan
from the aqueous solution (Agrafioti et al., 2013). The maximum et al., 2014; Cha et al., 2016; Braghiroli et al., 2018).
adsorption capacity of Ni2þ from water by chicken manure mixed
with sawdust-derived biochars was 11 mg/g at 650  C 3.3. Nitrogen and phosphorus removal
(Higashikawa et al., 2016). Marine macro-algae magnetic biochars
are rich in oxygen-functional groups, which attributes to their high Biochar can also absorb nutrients, such as nitrogen and phos-
selectivity and adsorption capacity to Cu2þ (69.37 mg/g for kelp phorus in aqueous phase (Zhang et al., 2012a, 2014; Yao et al.,
magnetic biochar and 63.52 mg/g for hijikia magnetic biochar) (Son 2013b; Zhang and Gao, 2013; Xue et al., 2016). Ammonium, ni-
et al., 2018). trate and phosphate are the common forms of reactive nitrogen and
phosphorus in wastewater, and can lead to eutrophication (Yao
3.2. Organic contaminant removal et al., 2012b; Yang et al., 2017; Xu et al., 2018). Table 4 lists the
adsorptions of nitrogen and phosphorus on various biochars in
Organic contaminants are another major type of pollutants in aqueous phase. The adsorption capacity of modified biochars for
aquatic environment, which include pesticides, herbicides, and nitrogen and phosphorus is significantly higher than pristine bio-
antibiotics etc.. Table 3 summarizes biochar adsorption of some chars, because the modified biochars have higher SSA, more reac-
organic contaminants in aqueous phase. Organic pollutants are tion activity and SFG.
toxic and can reduce dissolved oxygen in water and cause harm to Post-treatment of biochars have significant effects on ammo-
the aquatic ecosystem and human health (Ahmed et al., 2016). nium adsorption. Oxidized maple wood biochar has higher
Switchgrass biochar (SGB) and magnetic switchgrass biochar ammonium adsorption capacity than maple wood biochar (Wang
(MSGB) were employed to remove metribuzin herbicide from et al., 2016a). Additionally, pyrolysis temperatures affect ammo-
aqueous solutions. The low solution pH value is beneficial to bio- nium adsorption. Biochars produced from pine sawdust at 300  C
char for the metribuzin adsorption compared to the high solution shows the highest NHþ 4 adsorption capacity based on the higher H/
pH value. Metribuzin adsorption onto both SGB and MSGB is un- C and O/C ratios and presence of more functional groups on the
affected by temperature increase (Essandoh et al., 2017). Biochars surface of it (Yang et al., 2017). This study demonstrates that
can also remove antibiotics, such as sulfonamides and tetracyclines chemical bonding and polar interaction between NHþ 4 and SFG are
(Yao et al., 2012a; Sun et al., 2018). The mechanism underlying the likely mechanisms for enhanced NHþ 4 adsorption.
removal of sulfonamides and tetracyclines is probably due to the Pre-treatment of feedstock show pronounced effects on
electron donor-acceptor interactions and associated with the adsorption of phosphorus. The digested sugar beet tailing biochar
attracting groups on surface area rings (Peiris et al., 2017). Sulfa- shows the highest phosphate removal ability with a removal rate
methoxazole (SMX) is one of the typical sulfonamide antibiotics around 73% (Yao et al., 2011a). This is probably because the large
widely used for both human and animals. SMX adsorption onto the amount of colloidal and nano-sized periclase on its surface, which
digested bagasse biochars is mainly controlled by p-p interaction has a strong ability to bind phosphate in aqueous solution. Pre-
and effected by the solution pH value (Yao et al., 2018). Iron and zinc treatment can be performed during plant growth. For example,
doped sawdust biochar shows high simultaneous removal of the biochar derived from tomato plants that enriched with Mg
tetracycline from aqueous solution. The predominant adsorption during their growth, which shows increased adsorption of phos-
mechanisms include site recognition, bridge enhancement, and site phate in aqueous solution, reaching more than 100 mg/g (Yao et al.,
competition (Zhou et al., 2017b). 2013b). Additionally, biochars produced from wood waste pre-
In addition, several studies have also suggested biochar’s ap- treated with magnesium oxides (Mg-biochar) was used to recover
plications for adsorption of organic matter for water treatment, and ammonium and phosphate (Xu et al., 2018). The struvite
8 W. Xiang et al. / Chemosphere 252 (2020) 126539

Table 3
Biochar adsorption of organic contaminants in aqueous solutions.

Biochar Treatment/ Pyrolysis Biochar Organic Initial Adsorption Removal mechanism Ref.
feedstock Modification temperature dose (g/ contaminants concentration capacity (mg/
( C) L) (mg/L) g)

Switchgrass Magnetization 425 1 Metribuzin 100 39.6 Electrostatic attraction and hydrogen bonds Essandoh
herbicide et al.,
(2017)
Switchgras Pristine 425 1 Metribuzin 100 38.2 Electrostatic attraction and hydrogen bonds Essandoh
herbicide et al.,
(2017)
Bagasse Anaerobically 600 2 Sulfamethoxazole 10 1.6 p-p EDA interaction Yao et al.,
digested (2017)
Bagasse Anaerobically 600 2 Sulfapyridine 10 3.2 p-p EDA interaction Yao et al.,
digested (2017)
Bamboo Graphene 600 1 Sulfamethazine 10 6.5 p-p EDA interaction, pore-filling, cation exchange, Huang
sawdust oxide-coated hydrogen bonding interaction and electrostatic et al.,
interaction (2017)
Bamboo Pristine 600 1 Sulfamethazine 10 3.1 p-p EDA interaction, pore-filling, cation exchange, Huang
sawdust hydrogen bonding interaction and electrostatic et al.,
interaction (2017)
Sawdust Iron and zinc 600 / Tetracycline 150 86 Site recognition, bridge enhancement, and site Zhou et al.
doped competition (2017b)
Sawdust Iron and zinc 600 / Tetracycline 100 53.8 Site recognition, bridge enhancement, and site Zhou et al.
doped competition (2017b)
Peanut Magnetization 800 2 Trichloroethylene 9.2 4.6 Hydrophobic partitioning, pore-filling and reductive Liu et al.
shell degradation. (2019b)
Reed Magnetization 600 0.5 Florfenicol 20 5.3 Hydrogen bonding, pore-filling effect and p-p EDA Zhao and
interaction Lang,
(2018)
Reed Pristine 600 0.5 Florfenicol 20 2.6 Pore-filling effect and p-p EDA interaction Zhao and
Lang,
(2018)
Crab shell calcium-rich 800 1 Chlortetracycline 100 70 Cation bridging, electrostatic interaction, hydrogen Xu et al.,
biomass hydrochloride bonding and p-p interaction (2020)
Crab shell calcium-rich 800 1 Chlortetracycline 2000 1975 Adsorption and flocculation Xu et al.,
biomass hydrochloride (2020)

precipitation on the surface of biochar is the dominant mechanism wastewater. Biochars have been applied in the treatment of in-
for the removing ammonium and phosphate. Other reports have dustrial wastewater.
also shown modified biochars for removing the nitrate (NO 3 ), total A biochar mixed with chitosan after cross linking can be casted
Kjeldahl nitrogen (TKN), total nitrogen (TN), total phosphates (TP), into membranes, beads, and solutions. It can be effectively utilized
and phosphate (PO34 ) from aqueous solutions (Mohan et al., 2014; as an adsorbent for heavy metals adsorption in industrial waste-
Usman et al., 2016; Sun et al., 2017; Vikrant et al., 2017). A general water. The ratio of biochar and chitosan would affect the adsorption
conclusion is that the modifications change biochar surface of copper, lead, arsenic, cadmium and other heavy metals in in-
chemistry, thus resulting in enhanced nutrients sorption capacity dustrial wastewater (Hussain et al., 2017). Gliricidia biochar is a
compared with pristine biochars. promising material for crystal violet (CV) removal from an aqueous
environment in dye-based industries. The CV sorption process is
4. Biochar technology in wastewater treatment governed by the pH value, surface area and pore volume of biochar
(Wathukarage et al., 2017). Bagasse biochar was used to adsorb lead
As discussed above, biochars are effective adsorbents for from the battery manufacturing industry effluent. The maximum
removal of various contaminants due to its special properties, such adsorption capacity can reach 12.7 mg/g and the adsorptive process
as large SSA and abundant SFG. Thus, biochars have become is related to medium pH value, contact time and dosage (Poonam
increasingly important as a solution to remediate pollutants in the and Kumar, 2018). Biochar was also used to recapture nutrients
industrial and agricultural sectors for improving environmental from ammonium and phosphate-based dairy wastewater. Biochar
quality (Wang et al., 2017a). Wastewater has been a global issue, can adsorb 20e43% of ammonium and 19e65% of phosphate in
which is a byproduct of domestic, industrial, commercial or agri- flushed dairy manure within 24 h (Ghezzehei et al., 2014). Thus far,
cultural activities. Biochars have great potential to be used for most of the experiments on biochar application in removal of
wastewater treatment. This section mainly focuses on discussing contaminants from industrial wastewater were conducted in lab-
biochar’s applications in treatment of industrial wastewater, oratory setting, further research and implementation in real-world
municipal wastewater, agricultural wastewater and stormwater conditions is needed.
(Fig. 4).
4.2. Municipal wastewater treatment
4.1. Industrial wastewater treatment
Biochar can be directly used or combined with biofilter and
The industrial wastewater comes from various sources including other technologies for municipal wastewater treatment, which
mining, smelting, battery manufacturing, chemical industry, result in recovery of labile nitrogen and phosphorus (Cole et al.,
leather manufacturing, dyes, and others. And the pollutants are 2017). Engineered biochar loaded with aluminum oxyhydroxides
mainly heavy metals and organic pollutants in industrial (AlOOH) was applied to recycle and reuse phosphorus from
 W. Xiang et al. / Chemosphere 252 (2020) 126539 9

Table 4
Biochar adsorption of nitrogen and phosphorus in aqueous solutions.

Biochar feedstock Treatment/ Pyrolysis Biochar Nutrient Initial Adsorption Removal mechanism Ref.
Modification temperature dose (g/ concentration capacity (mg/
( C) L) (mg/L) g)

Pine sawdust Pristine 300 3 NHþ

4 100 5.38 Chemical bonding and electrostatic Yang et al.,
interaction of NHþ4 with the surface functional (2017)
groups.
Wheat straw Pristine 550 3 NHþ
4 100 2.08 Chemical bonding and electrostatic Yang et al.,
interaction of NHþ4 with the surface functional (2017)
groups.
Wood waste MgO modified 600 2 NHþ
4 8203 47.5 Struvite precipitation Xu et al.,
(2018)
Sugarcane harvest MgO particle- 550 1.25 NHþ
4 200 22 Struvite crystallization, electrostatic Li et al.,
residue impregnated attraction, and p-p interactions (2017)
Wheat straw MgeFe layered 600 2 NO
3 45 24.8 Surface adsorption and interlayer anion Xue et al.,
double hydroxides exchange (2016)
(LDH)
Peanut shells MgCl2 solution 600 2 NO
3 20 94 Surface adsorption Zhang et al.
immersed (2012a)
Hickory wood chips Aluminum salt 600 2.5 Phosphorus 6.4 8.346 Electrostatic attraction Zheng et al.
treated (2019a)
Wheat straw Acid wash and 500e560 12.5 Phosphorus 25 1.06 Adsorption and surface precipitation Dugdug
water wash et al.,
(2018)
Hardwood Acid wash and 500e550 12.5 Phosphorus 25 1.2 Adsorption and surface precipitation Dugdug
water wash et al.,
(2018)
Willow wood Acid wash and 500e550 12.5 Phosphorus 25 1.93 Adsorption and surface precipitation Dugdug
water wash et al.,
(2018)
Wood waste MgO modified 600 2 PO3-
4 318.5 116.4 Struvite precipitation, surface adsorption Xu et al.,
(2018)
Bamboo MgeAl layered 600 2 PO3-
4 50 13.11 Interlayer anion exchange and surface Wan et al.,
double hydroxides adsorption (2017)
(LDH)
Anaerobically Pristine 600 2 PO3-
4 61.5 25 Surface adsorption by colloidal and nano- Yao et al.
digested sugar sized MgO particles (2011b)
beet tailings
Cottonwood AlCl3 solution 600 2 PO3-
4 1600 135 Adsorption by unique nanostructure Zhang and
immersed Gao,
(2013)
Sugar beet tailings MgCl2 solution 600 2 PO3-
4 1600 835 Surface adsorption Zhang et al.
immersed (2012a)
Tomato leaves Mg enriched 600 2 PO3-
4 588.1 100 Precipitation, surface deposition Yao et al.
(2013a)
Cottonwood HTC þ LDH 180 2 PO3-
4 2000 386 Surface adsorption Zhang
et al.,
(2014)

secondary treated wastewater (Zheng et al., 2019a). The adsorption Wastewater from residential units not connected to any municipal
mechanism of phosphorus is mainly through electrostatic attrac- sewage treatment plant was treated with biochar in on-site sewage
tion. Phosphorus adsorbed on engineered biochar can be utilized as treatment facility (OSSFs) (Blum et al., 2018). The addition of bio-
a slow-release fertilizer for crop production. char obviously increases the removal rate of some polar and hy-
Biochar produced from digested sludge was used as an adsor- drophilic compounds. OSSFs thus can be upgraded with low-cost
bent for ammonium removal from municipal wastewater. Biochar biochar adsorbents.
derived at 450  C has the highest ammonium removal capacity
attribute to its higher surface area and functional group density,
and the process is controlled by chemisorption (Tang et al., 2019). 4.3. Agricultural wastewater treatment
Biochar derived from waste sludge was used as catalysts to ozonate
refinery wastewater and shows high removal rate of the total Agricultural contamination is becoming increasingly serious
organic carbon. Because the biochar contains functional carbon due to the rapid development of agricultural industry, more and
groups, Si/O structures, and metallic oxides, it can promote oxida- more pesticides or toxic heavy metals are discharged into farm-
tion through the formation of hydroxyl radicals and mineralized lands (Wei et al., 2018). Many researchers have applied biochar and
petroleum contaminants (Chen et al., 2019). its modified forms to treatment of agricultural wastewater
Municipal wastewater can be treated with biochar, produced contamination.
from municipal biowaste, at the biofiltration stage. Biochar has a Pentachlorophenol and atrazine are two most common pesti-
high porous surface area that allows it to act as a biofilter in cides in agriculture. Rice straw biochar and phosphoric acid
municipal wastewater treatment. The COD, TSS, TKN and TP of modified rice straw biochars show significantly high adsorption for
wastewater reduce 90%, 89%, 64%, and 78%, respectively, after being imidacloprid and atrazine from agricultural wastewater (Mandal
passed through the biochar biofilter (Manyuchi et al., 2018). and Singh, 2017). Soybean and corn straw biochar both show
high atrazine removals and the adsorption capacities are mainly
10 W. Xiang et al. / Chemosphere 252 (2020) 126539

Fig. 4. Biochar application in wastewater treatment.

due to the pore volume and pH value of biochar (Zhao et al., 2013; adsorption capacities are closely correlated with nano-material
Liu et al., 2015). Steam-activated biochar can effectively remove content, SSA, SFG, and porous structures (Cha et al., 2016;
sulfamethazine and the removal rate is pH value dependent Braghiroli et al., 2018; Son et al., 2018; Wan et al., 2018; Yao et al.,
(Rajapaksha et al., 2015). Zero valent iron magnetic paper mill 2018). In addition, the adsorption mechanism by biochars are
sludge biochar (ZVI-MBC) was used for removal of pentachloro- affected by inner-sphere complexes, p-p interaction, hydrophobic
phenol (PCP) from the effluent (Devi and Saroha, 2014). The ZVI- effect, precipitation, ion exchange, and so on (Yuan et al., 2011; Cha
MBC can simultaneously adsorb and dechlorinate the PCP in the vre et al., 2018; Wei et al., 2018; Yao et al., 2018).
et al., 2016; Lefe
effluent and achieve the complete removal of PCP. The removal of
glyphosate, diuron and carbaryl from agricultural wastewater by
biochar have been also investigated. The adsorption capacity of 4.4. Stormwater treatment
biochar to pesticides are related to biochar feedstock, functional
materials, and target contaminants (Wei et al., 2018). With the development of urbanization, urban stormwater
The toxic heavy metals in agricultural wastewater is another runoff has been widely concerned due to its influence on water
pervasive problem. The common concerned toxic metals include quality. Stormwater runoff can significantly contribute to the
As, Cr, Cu and Pb (Table 2). The adsorption capacity of Cu2þ and As5þ degradation of natural water quality and requires treatment before
in agricultural wastewater by biochar can reach 69.4 mg/g and discharge, which is mainly due to increased concentrations of
34.1 mg/g, respectively; and the adsorption quantity of Cd2þ and metals, organic matter and biological pollutants (Mohanty et al.,
Pb2þ are ranged from 0.4 mg/g to 12.3 mg/g, and 36 mg/g to 35 mg/ 2014; Gray, 2016; Tian et al., 2016; Ulrich et al., 2017; Ashoori
g, respectively (Higashikawa et al., 2016; Cho et al., 2017; Zhou et al., 2019).
et al., 2017a; Son et al., 2018). For the heavy metals in agricultural Bioretention and biofiltration are commonly used for storm-
wastewater, the possible adsorption mechanisms usually involve water treatments, but the purification of stormwater contaminants
electrostatic interactions, surface complexation, ion exchange, by these two systems is not ideal (Gray, 2016; Lau et al., 2016; Ulrich
intermolecular interaction, cation-p bonding, and p-p interactions et al., 2017). Biochar and its modified forms, as the effective media,
(Wei et al., 2018). have been applied to stormwater treatment systems (Fig. 5). A
The adsorption behavior of biochars for various agricultural recent study shows that an aluminum-impregnated biochar can
contaminants differs widely (Wei et al., 2018). In general, the effectively remove As5þ and other runoff pollutants, such as Pb2þ,
Zn2þ, Cu2þ, and PO3 4 , in a polluted urban water runoff (Liu et al.,
 W. Xiang et al. / Chemosphere 252 (2020) 126539 11

Fig. 5. Biochar application in stormwater treatment: (a) Potential functions of biochar at different region of bioinfiltration system (Mohanty et al., 2018). (b) Schematic diagram of
the enhanced stormwater contaminants removal by biochar-amended biofilters (Lu and Chen, 2018).

2019a). A biochar-based filtration medium has been effectively In general, biochar has been used as filter media in stormwater
deployed to remove copper and zinc in stormwater runoff, and the treatment. Various removal capacities of contaminants in storm-
remove rate reached more than 85% and 95%, respectively. But the water depend on biochar properties, pollutant characteristics, and
biochar filtration media need to be carefully tested and designed to aqueous chemistry (Mohanty et al., 2018). Biochar is more feasible
meet the requirements of stormwater treatment (Gray, 2016). and promising than other materials used in stormwater treatment,
Biochars have been integrated with biofilters for removing because it is inexpensive and readily available and has many
bisphenol A (BPA) from stormwater. Wood dust biochar shows a beneficial functions in stormwater treatment systems.
high adsorption efficiency and increased capacity of BPA attribute
to its high SSA and pore volume, which also promotes phragmites
australis growth, increases E. coli, TOC, TSS, nitrogen and phos- 5. Conclusions and future perspectives
phorus removal rates (Ashoori et al., 2019). Biochar amendment has
improved the removal of contaminant in stormwater biofilters, Biochar is an efficient and low-cost adsorbent, which can be
particularly the toxic trace organic contaminants (TOrCs) that have produced from a variety of biomass materials including agricultural
been poorly removed in conventional systems. Biochar-amended crop residues, forestry residues, sewage sludge, manures, solid
biofilter columns can maintain more than 99% TOrC removal rate organic municipal wastes, and thus has been used in wastewater
compared to the unamended biofilter columns. Meanwhile, treatment. This article reviews the current technologies for biochar
biochar-amended biofilter can increase the removal of TOC, TN, and production with an emphasis on feedstock pre-treatment, thermal
TP greater than 60% (Ulrich et al., 2017). conversion, and post treatment technologies. It summarizes the
Poultry litter biochars (PLB) pyrolyzed at 500  C were applied to biochar application in wastewater treatment including industrial
adsorb ammonium in stormwater treatment systems. There is a wastewater, municipal wastewater, agricultural wastewater and
significant positive correlation between NHþ 4 sorption and biochar
stormwater. Mechanisms underlying the biochar adsorption of
CEC. The ion competition in stormwater adsorption experiments contaminants are discussed.
suggests that NHþ 4 adsorption is dominated by cation exchange
The main conclusions of this review are as follows: (1) Biochar
(Tian et al., 2016). Zn-activated sewage sludge-based activated properties are related to the type of feedstock, feedstock pre-
carbon can remove PO4eP and NO3eN effectively from leachate treatment technology, thermal process, and post-treatment of
made from stormwater. And the removal rates of PO4eP and biochars. The modifications of biochars by increasing the SSA, re-
NO3eN decrease with increasing pH value (Yue et al., 2018). Biochar action activity or by forming functional groups, become increas-
and zero valent iron (ZVI) amending bioretention cells can increase ingly important as a new and exciting area of engineered biochar
the NO-3 removal performance in stormwater system, which pro- research and its application for improving environmental quality.
vides an important prospect for increasing nitrate removal effi- (2) Largely due to the modifications, engineered biochar as an
ciency in bioretention systems (Tian et al., 2019). adsorbent to remove aqueous contaminant, such as heavy metals,
Biofilters/bioretention system with biochar can also effectively organic contaminants, nitrogen and phosphorus is controlled by
remove microorganisms from stormwater (Mohanty et al., 2014; various mechanisms, mainly including ion exchange, adsorption,
Lau et al., 2016). Biofilters amended with 5% biochar can retain up surface precipitation, surface complexation etc. (3) The potential of
to 3 orders of magnitude more E. coli, and prevent their mobiliza- biochar for removal of pollutants from industrial wastewater,
tion during successive intermittent flows. This indicates that municipal sewage, agricultural sewage, and stormwater has been
amending biofilters with biochar can improved the removal of well demonstrated in laboratory. Its application for onsite appli-
bacteria from stormwater (Mohanty et al., 2014). H2SO4-modified cation requires further investigation. Although number of re-
wood biochar can be an effective bioretention filter medium for searches have been done on production and application of biochar
E. coli removal from stormwater. It improves E. coli retention and in wastewater treatment, there are still knowledge gaps that need
reduces remobilization. The results indicate that the transport of to be filled.
E. coli is governed by the morphology structures and hydropho- Additional studies are still need to: (1) develop the new low-cost
bicity of the biochars (Lau et al., 2016). and high-efficiency modification technology of biochar, (2) increase
the practical application of biochar in wastewater treatment,
12 W. Xiang et al. / Chemosphere 252 (2020) 126539

especially in industrial wastewater and municipal wastewater an adsorbent for the removal of pentachlorophenol from the effluent. Bioresour.
Technol. 169, 525e531.
treatment, and (3) further improve the adsorption capacity of
Dugdug, A.A., Chang, S.X., Ok, Y.S., Rajapaksha, A.U., Anyia, A., 2018. Phosphorus
biochar on heavy metals, organic contaminants, nitrogen and sorption capacity of biochars varies with biochar type and salinity level. Envi-
phosphorus. ron. Sci. Pollut. Res. Int. 25, 25799e25812.
Dura n-Jime nez, G., Herna ndez-Montoya, V., Montes-Mor an, M.A., Kingman, S.W.,
Monti, T., Binner, E.R., 2018. Microwave pyrolysis of pecan nut shell and ther-
Declaration of competing interest mogravimetric, textural and spectroscopic characterization of carbonaceous
products. J. Anal. Appl. Pyrol. 135, 160e168.
Essandoh, M., Wolgemuth, D., Pittman, C.U., Mohan, D., Mlsna, T., 2017. Adsorption
The authors declare that they have no known competing
of metribuzin from aqueous solution using magnetic and nonmagnetic sus-
financial interests or personal relationships that could have tainable low-cost biochar adsorbents. Environ. Sci. Pollut. Control Ser. 24,
appeared to influence the work reported in this paper. 4577e4590.
Fang, J., Zhan, L., Ok, Y.S., Gao, B., 2018. Minireview of potential applications of
hydrochar derived from hydrothermal carbonization of biomass. J. Ind. Eng.
Acknowledgements Chem. 57, 15e21.
Ghezzehei, T.A., Sarkhot, D.V., Berhe, A.A., 2014. Biochar can be used to capture
essential nutrients from dairy wastewater and improve soil physico-chemical
W.X and X.Z. would like to acknowledge the support of the properties. Solid Earth 5, 953e962.
Natural Science Foundation of the Jiangsu Higher Education In- Gonz alez, M.E., Cea, M., Reyes, D., Romero-Hermoso, L., Hidalgo, P., Meier, S.,
stitutions of China (Grant No. 18KJA610003), Key R & D Projects of Benito, N., Navia, R., 2017. Functionalization of biochar derived from lignocel-
lulosic biomass using microwave technology for catalytic application in bio-
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from aqueous solutions: mechanisms and potential application as a slow- Zhang, X., Xiang, W., Wang, B., Fang, J., Zou, W., He, F., Li, Y., Tsang, D.C.W., Ok, Y.S.,
release fertilizer. Environ. Sci. Technol. 47, 8700e8708. Gao, B., 2019a. Adsorption of acetone and cyclohexane onto CO2 activated
Yao, Y., Gao, B., Chen, J.J., Zhang, M., Inyang, M., Li, Y.C., Alva, A., Yang, L.Y., 2013b. hydrochars. Chemosphere 245, 125664.
Engineered carbon (biochar) prepared by direct pyrolysis of Mg-accumulated Zhang, Y., Chen, P., Liu, S., Peng, P., Min, M., Cheng, Y., Anderson, E., Zhou, N., Fan, L.,
tomato tissues: characterization and phosphate removal potential. Bioresour. Liu, C., Chen, G., Liu, Y., Lei, H., Li, B., Ruan, R., 2017b. Effects of feedstock
Technol. 138, 8e13. characteristics on microwave-assisted pyrolysis - a review. Bioresour. Technol.
Yao, Y., Gao, B., Fang, J., Zhang, M., Chen, H., Zhou, Y., Creamer, A., Sun, Y., Yang, L., 230, 143e151.
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Yao, Y., Gao, B., Inyang, M., Zimmerman, A.R., Cao, X., Pullammanappallil, P., Yang, L., Zhao, H., Lang, Y., 2018. Adsorption behaviors and mechanisms of florfenicol by
2011a. Biochar derived from anaerobically digested sugar beet tailings: char- magnetic functionalized biochar and reed biochar. Journal of the Taiwan
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6273e6278. Zhao, T., Yao, Y., Li, D., Wu, F., Zhang, C., Gao, B., 2018. Facile low-temperature one-
Yao, Y., Gao, B., Inyang, M., Zimmerman, A.R., Cao, X.D., Pullammanappallil, P., step synthesis of pomelo peel biochar under air atmosphere and its adsorption
Yang, L.Y., 2011b. Removal of phosphate from aqueous solution by biochar behaviors for Ag(I) and Pb(II). Sci. Total Environ. 640e641, 73e79.
derived from anaerobically digested sugar beet tailings. J. Hazard Mater. 190, Zhao, X., Ouyang, W., Hao, F., Lin, C., Wang, F., Han, S., Geng, X., 2013. Properties
501e507. comparison of biochars from corn straw with different pretreatment and
Yao, Y., Gao, B., Wu, F., Zhang, C., Yang, L., 2015. Engineered biochar from biofuel sorption behaviour of atrazine. Bioresour. Technol. 147, 338e344.
residue: characterization and its silver removal potential. ACS Appl. Mater. In- Zheng, Y., Wang, B., Wester, A.E., Chen, J., He, F., Chen, H., Gao, B., 2019a. Reclaiming
terfaces 7, 10634e10640. phosphorus from secondary treated municipal wastewater with engineered
Yao, Y., Gao, B., Zhang, M., Inyang, M., Zimmerman, A.R., 2012b. Effect of biochar biochar. Chem. Eng. J. 362, 460e468.
amendment on sorption and leaching of nitrate, ammonium, and phosphate in Zheng, Y., Yang, Y., Zhang, Y., Zou, W., Luo, Y., Dong, L., Gao, B., 2019b. Facile one-step
a sandy soil. Chemosphere 89, 1467e1471. synthesis of graphitic carbon nitride-modified biochar for the removal of
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and sulfapyridine (SPY) from aqueous solutions by biochars derived from 89e96.
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Yao, Y., Zhang, Y., Gao, B., Chen, R.J., Wu, F., 2018. Removal of sulfamethoxazole excellent Pb(II) adsorption property produced from fresh and dehydrated ba-
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biochar stabilization of metal-contaminated soils. Sci. Total Environ. 616.
 Identifying the Major Sources of
Nutrient Water Pollution
A national watershed-based analysis connects nonpoint
and point sources of nitrogen and phosphorus with
regional land use and other factors.
L A R R Y 1. P U C K E T T

ince the passage of the Clean Water Act fective plans, it is necessary to identify and quan-
(CWA] in 1972, more than $540 billion has tify the dominant sources of nonpoint-source
heen spent on water pollution controls. pollution at the watershed scale, and then elimi-
In spite of this massive commitment of nate, to the greatest extent possible, those sources
funds, in 1992 approximately44% of US. contributing most to water quality problems.
river miles tested still did not fully sup- Using information from various databases, such
port the uses designated by the states (11. One well- as the U.S. Geological Survey’s National Water Qual-
recognized problem is the lack of controls in the CWA ity Assessment program, Census of Agriculture, and
for nonpoint-source pollution (21. the National Trends Network, it is now possible to
Preventing or controlling pollution from non- estimate the relative magnitudes of the major non-
point sources has been stymied, in part, because lit- point and point sources of nutrients. The estimates
tle was known about the relative magnitudes of non- allow three key points to be addressed the propor-
point and point sources of nutrients at the national tion of nonpoint and point sources of nutrients in
level. Only recently has it heen possible to estimate US. watersheds, the continuing importance of point
the major sources of nutrients entering watersheds sources of nutrients, and the relative importance of
and relate these inputs and resulting stream loads atmospheric inputs as a nonpoint source of nitro-
to land use patterns and regional settings. gen.
Analysis of these new data reveals that nutrient
inputs into watersheds vary according to land use Identifying the major sources
practices, that atmospheric nitrogen may he a more Among the various sources of water quality impair-
important source of nutrient contamination than pre- ment, agricultural practices are ranked as the most
viously believed, and that in some localities point impoaant factor for rivers and lakes and third in im-
sources are still the major water quality problem. portance for estuaries (1). Going one step further and
These findings underscore the need for individual- delineating agriculturalpractices with respect to spe-
ized watershed management plans for preventing and cific pollutants, nutrient contamination is identi-
controlling water pollution in the United States. fied as the second most important contribution to
The CWA has traditionally focused on reducing river and lake pollution and first in importance for
discharges of pollutants, particularly nitrogen and estuaries.
phosphorus, to surface waters from sewage treat- Commercial fertilizer is the primary agricultural
ment plants and other point sources. Because of the nonpoint source of nitrogen and phosphorus. Be-
Act’s point-source focus, nearly 90% of the monies tween 1945 and 1993, the use of nitrogen in com-
allocated to water pollution prevention have been mercial fertilizers in the United States increased 20-
targeted at point sources I J ) . Congress has been con- fold, rising from about 0.5 million to nearly 10.3
sidering revisions to the Act since 1993, but it is un- million metric tons per year. Phosphorus use has in-
clear at this time how nonpoint-source pollution will creased as well, kom about 0.5 million to nearly 1.8
be addressed. million metric tons per year [Figure 1).
In spite of Congressional uncertainty, many state Much of the fertilizer is applied in the upper Mid-
and federal water quality programs have begun to west “corn belt” states (Figure 21. Studies have es-
tackle nonpoint sources by focusing on watershed- timated that farmers may apply 24 to 38% more fer-
based management plans. However, to develop ef- tilizers than crops require because of uncertainties

408 A .
VOL. 29, NO. 9.1995 iENVIRONMENTAL SCIENCE &TECHNOLOGY 0013-936W95/0929-408A109.00iO B 1995 American Chemical Society
, ... .,,..

Agricultural practices are the moa important factor in water quality impairment for rivers and lakes (above). but in lage
urban centers point sources remain a major nutrient source (below).

associated with weather and soil nutrient status (3.4). nure is not accounted for when computing crop re-
Animal manure is another agricultural pollution quirements, the excess may move into surface and
source. Each year in the United States the manure ground waters.
from 7.5 billion farm animals produces an esti- Atmospheric inputs have been largely ignored as
mated 5.9 million and 1.8 million metric tons of ni- nonpoint-source pollution because they originate pri-
trogen and phosphorus, respectively The upper Mid- marily at point sources. For example, releases of ni-
west has the greatest manure nutrient input (Figure trogen oxides into the air from point sources be-
3).Where farm animals are allowed to graze freely, come nonpoint sources ofwater pollution when that
the manure is distributed over the landscape and rep- nitrogen reaches water bodies as rainfall. More than
resents a nonpoint source of nutrients. However, 2.9 million metric tons of nitrogen are deposited in
where animals are confined to feedlots, barns, or the United States each year from the atmosphere (3,
sheds, they may become more of a point source nu- primarily in the northeastern states (Figure 4). About
trient problem. lfthe nutrient content of applied ma- 53%of these nitrogen emissions come from coal- and

VOL. 29. NO. 9.1995 I ENVIRONMENTAL SCIENCE &TECHNOLOGY. 409 A

 have been developed. The data were examined for
differences and similaities according to land use clas-
Changes in fertilizer nitrogen and phosphorus use $ificationand geographical region; differences that
U S data show that from 1945 to 1993 nrtrogen use jumped 20-fold and ire cited are statisticallysignificant. Land use classes
ihorus use tripled Table 1) identify the dominant use of land within
he watershed hut do not eliminate other uses that
can affect water quality
Fertilizer nitrogen inputs per unit area to agri-
cultural land (Table 2) were largest in watersheds
dominated by agriculture, Less in forested water-
sheds, and lowest in urban-dominated watersheds.
nputs of fertilizer in mixed watersheds were greater
than in urban watersheds, hut there was no statis-
tically significant difference from inputs in agricul-
ural and forested watersheds,
There were no statistically significant differ-
ences among phosphorus inputs in agricultural, for-
ested, and mixed-land-use watersheds. However, all
Yhowed significantly greater inputs than did urban
Natersheds.
The higher fertilizer inputs of both nitrogen and
3hosphorus per unit area to agricultural lands in ag-
ricultural- versus urban-dominated watersheds sug-
gest a systematic difference in agricultural prac-
tices as well as higher application rates in
predominantly agricultural watersheds, This could
also mean that more of the agricultural land in ag-
Watershed class tion by dominant land u ricultural-dominated watersheds is used for inten-
sive crop production. Conversely, the significantlylow
Values under the land use classifications are the percent; ~’~ ~

fertilizer application rates in urhan-dominated wa-

each category in the watershed, except for populi e tersheds suggest less intensive use of agricultural
which is persons per square kilometr- lands or possibly conversion to other uses. The higher
inputs of nitrogen and phosphorus m forested and
Dominant land Cmp and ropuit mixed watersheds relative to urban watersheds re-
U o C C l ~ partun ~ ~ Form urban dsns tlect the importance of agricultural land as a source
Agricultur 40 C40 < 10 of nutrients in those watersheds.
Urban 30 >3 Nitrogen inputs to agricultural land from fertil-
Forest 40 > 50 < 10 izer (Table 2) were largest in western and southeast-
Mixed watersheds not meeting the above crit ern watersheds and lowest in northeastern water-
sheds. Inputs in the central watersheds were not
statistically significantlydifferent from the other wa-
tersheds.
oil-burning electric utilities and large industries (6); Phosphorus inputs to agricultural land from fer-
approximately 38% from automobiles, trucks, buses, tilizer were significantly larger in western than in cen-
and other forms of transportation; and the remain- tral watersheds. Inputs in southeastern and north-
ing 9% from miscellaneous sources. eastern watersheds were not statisticallysignificantly
Ammonium can also he an important form of ni- different from western and central watersheds.
trogen in precipitation. However, its predominant The consistently large nitrogen and phosphorus
sources are volatilization from fertilizer and animal fertilizer inputs in western watersheds reflect the in-
manure. tensive agriculturalpractices in areas like the San Joa-
Overall, the most recent nationwide estimates of quin Valley CA (Figure 21, where multiple crops in
point sources of pollution [compiled in 1984) indi- the same year may require multiple fertilizer appli-
cated that from 1978 to 1981 point sources dis- cations.
charged approximately 1.2 million metric tons of ni- There were no significant differences in nitro-
trogen and 260,000 metric tons of phosphorus gen or phosphorus inputs to agricultural land from
annually [7).Thesevalues are dwarfed by the 19.4 mil- animal manure based on land use (Table 21. On a re-
lion tons of nitrogen and 5.7 million metric tons of gional basis nitrogen and phosphorus inputs were
phosphorus from nonpoint sources. Yet, in spite of largest in southeastern and northeastern water-
the billions of dollars spent on point source pollu- sheds and lowest in central watersheds. Inputs in
tion controls, point sources may still represent a sig- western watersheds were not statistically signifi-
nificant local impact (I). cantly different from those in other regions [Figure
3).
Agricultural practices and regional differences The large manure nutrient inputs to agricultural
Using information from various databases (see side- land in the northeast reflect the higher density an-
bar, Data sources and methodology), national esti- imal production in that region, particularly for the
mates based on nutrient inputs and stream loads dairy industry. Confined feeder operations for the pro-

4 1 0 A m VOL. 29. NO, s. 1995, ENVIRONMENTAL SCIENCE & TECHNOLOGY

 Median values for annual loadings
- to watersheds
Values followed by the same letter are not significantlydiflerent (a= 0.051 and apply only u1 multiple comparisons among land use
classes or regions listed in the rows. Retention is defned as the traction of total nnrogen or phosphorus inpuis not accounted for b
stream loads
Laad un clan Region
Elmad Apricubrel Ford Mixsd Urban Watcm Csml Souheartom Norheaalo
Atmosphere Nitrogen 0.35a 0.68 a,b 0.78 a,b 0.87 b 0.17 a 0.39 b 0.64 b 0.99 c
Fertilizer Nitrogen 6.06 a 2.88 b 3.24a.b 0.01 c 7.29 a 3.42 a,b 4.04a 2.30 b
Phosphorus 0.75 a 0.69a 0.76a 0.00 b 0.84 a 0.38 b 0.46 a,b 0.68 a.1
Manure Nitrogen 1.50 a 3.18a 2.10a 2.85 a 2.63 a,b 1.41 a 1.91 b 3.63 b
Phosphorus 0.65 a 0.73a 0.70a 0.53 a 0.63 a,b O37a 0.72 b 0.84 b
Point source Nitrogen 0.03 a,b 0.00 a 0.14 b,c 0.18~ 0.00 a 0.16 b 0.06 a,b 0.10 b
Phosphorus 0.01 a,b O.00a 0.02 b 0.05 b 0.00 a 0.03b 001 a.b 0.01 b
Stream load Nitrogen 0.61 a,b 0.31 a 0.59 b 0.39 a,b 0.30 a 0.30 a 0.33a 0.66 b
Phosphorus 0.03 a 0.02s 0.06 b 0.03 a,b 0.02 a 0.05 a 0.02 a 0.04 a
c
Retention Nnmgen 0.93 a 0.82 b 0.84a.b 0.77 b 0.86 a,b 0.90 a 0.89 a 0.76 b
Phosphorus 0.98 a 0.93 b 0.93a.b 0.86 b 0.96 a 0.93 a 0.95 a 0.93
s(
tl

duaion of pouliq and hogs also may account for pari

of this pattern.
rogen, phosphorus ferb'liier inputs
Atmospheric deposition and point sources cording to 1987 data, most fertilizer is applied in upper midweste
Atmospheric inputs of nitrogen (Table 2) were l a - tes. Inputs are in memc tons per square kilometer, by county.
est in predominantly urbanwatersheds and smali-
est in agricultural watersheds, whereas forested and trogon .essthan 0.35
mixed-land-usewatersheds were not statisticallysig- 1.35to 1.05 I 2.45to 10.5
nilicantly different from other types of land use. There
was a pronounced west-east trend of increasing ni-
trogen inputs, from the smallest inputs in western
watersheds to the largest inputs in noaheastern wa-
tersheds.
The large inputs in predominantly urban water-
sheds are due, in part, to large emissions of nitro-
gen oxides in urban areas. Also, many of the urban
watersheds are located in the northeast, where at-
mosphericinputs are larger (Figure 4). On the other
hand.. a -e r i c u l t d v dominated watersheds mav be
located at some distance from urban areas or in the
western pari of the nation, where atmospheric in-
puts are not as great. Atmospheric inputs in for-
ested and mixed-land-use watersheds fall between
the extremes of urban and agricultural watersheds,
because they ate scattered throughout the nation and
are influenced less by strong regional trends.
Nitrogen inputs from point sources (Table2) were
greatest in urban watersheds and smallest in for-
ested watersheds Values for agricultural and mixed-
land-use watersheds fell between urban and for-
ested watersheds. Phosphorus inputs followed
nitrogen trends, with the exception that phospho-
IUS & not significantlyp a t e ; in urban than & ag-
ricultural watersheds.
It was not surprising that point source nutrient
inputs were greatest in urban watersheds because
generally there are more point source dischargers in
urban areas. Mixed-land-use watersheds may incor-
porate large urban areas, accounting for their fre-
quently large point source inputs. Conversely, pre- puts from point sources (Table2) were largest in cen-
dominantly forested watersheds would be expected tral and northeastern watersheds, and smallest in
to have the smallest point source inputs because ma- western watersheds. The southeast inputs were not
jor point sources are less likely to be located there. significantly different from watersheds in other re-
On a regional basis nitrogen and phosphorus in- gions. The large inputs in the northeast result from

VOL.
2s. NO. s, 19% i ENVIRONMENTALSCIENCE &TECHNOLOGY 41 1 A
I - -
-.. ' a:..
3
.'
Nitrogen, phosptorus inputs from animal'manure
'
~
I
'

Data collected for 1987 showthatthe upper midwest has the greatest manure
nutrient input Inputs are in metric tons par square kilometer. by counoh.
,

' ' :
*,
the high numbers of urban watersheds, whereas the
large inputs in the central region are traced to mixed-
land-use watersheds. The small inputs in the west-
e m watersheds are due to the predominance of for-
ested watersheds in that sample population, and the

r ---
Nitrogen ILess than 0.35
00.35 to 0.70
I
I
0.70 to 1.4
1.40 to 12.
intermediate values in the southeast reflect the large
number of forested watersheds there.
Based on land use, nitrogen and phosphorus loads
(Table 2 ) followed similar patterns. The largest loads
were seen in mixed watersheds, the smallest in for-
ested watersheds; loads in @cultural and urban wa-
tersheds were not statistically significantly different
from others. As before, nitrogen loads were signifi-
cantly larger in the northeast than in other regions.
Phosphorus loads did not differ significantlyby re-
gion.
The large nitrogen and phosphorus loads in mixed
watersheds, and intermediate loads in agricultural
and urban watersheds, suggest that a mixture of non-
point and point sources of nutrients, as in the pre-
dominantly mixed-land-use watersheds, produced
the largest loads. The large nitrogen stream loads in
northeastern watersheds (as compared to the re-
sults for phosphorus. which did not vary signifi-
cantly by region) may result from a combination of
the significantly larger point-source and atmo-
spheric inputs of nitrogen in the northeast.
Retention of nitrogen and phosphorus (defined
as the fraction of total nutrient input from fertiliz-
ers, manure, atmospheric deposition, and point
sources not found in stream loads [see Table 21) was
greater in agriculturally dominated watersheds than
in the forest and urban watersheds. Mixed-land-
use watersheds were not statistically different. The
retention results for land use comparisons were con-
sistent with the assumption that more of the nutri-
ents from fertilizer, manure, or atmospheric depo-
sition would be retained by crops in the agricultural
watersheds. Lower retention may he seen in for-
ested watersheds hecause forests are located at higher
elevations where atmospheric inputs are greater, or
lands are steeper and rockier with shallower soils. Ur-
ban watersheds have large areas of pavement and
other impermeable materials that result in faster run-
off of atmospheric inputs. They also provide a greater
nitrogen potential for direct inputs of nutrients to streams from
nitrogen deposited from the point sources.
r square kilometer, by count\ Regionally, there were no significant differences
For phosphorus retention (Table 2). However, nitro-
ss than 0.32 - J.46 to 0.63 gen retention decreased from central and southeast-
32t00.46 IO.63tol.l6 ern watersheds, to western watersheds, to north-
eastern watersheds. The smaller retention values for
nitrogen in the northeast compared with the rest of
the watersheds are consistent with the finding of sig-
nificantly larger stream loads there as well. Other con-
tributing factors include the larger percentage of ur-
ban- and forest-dominated watersheds in the
northeast as well as the significantly larger atmo-
spheric inputs compared with the other regions.

Watershed-by-watershed variability
This study shows that different land use practices re-
5ult in variable inputs of nutrients to watersheds
around the nation. Major differences in the magni-
tudes of the various nonpoint and point sources of
nutrients are found even in neighboring water-

412 a VOL. 29. NO. 9.1995 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

 Data sources and methodology watershed and then multiplied by the area of the watershed to
estimate total inputs. Adlustments were made for droplet depo:
Watersheds were selected from among those investigated in the lion. urban effects, and dry deposition 16).
U.S. Geological Survey's (USGS) National Water Quality Assess- Point-source astmates were developed from data collectec
ment program. These program study units were first imple- oy E W s Permit Compliance System. Effluent volumes were mu
mented in 1991. tiplied by the reported concentrations of nitrogen and phospho
A modified version of the scheme employed by Smith and rJs for each facility located n the watersned to yield estimate!
others (8)was used to classify 114 watersheds based on the of me total mput to a given stream ( 1 0.
percent of various land use classes (agriculture, forest, urban, Mass transport of total n:trogen and total phosphorus n
mixed). This provided an indication of the land use in the water- streams was calculated for typical-flow years iusing the mini-
shed that would be most likely to have the greatest impact on mum variance Jnbiased estimator method 112. Mi witn conce
water quality (Table 1).Watersheds were also classified accord- tration oata available from USGS and the EPA databases.
ing to region (Western, Central, Southeastern, Northeastern). Retention was calculated as the fraction of the total nutrv
providing a means of grouping watersheds along broad hydro- ent mputs to the watershed from manure, fertilizer, air, and
geologic, climatic, and vegetation zones. po;nt soLrces that could not be accounted for n the stream
Estimates of nitrogen and phosphorus in commercial fertilizer loao. Althougn tnese four sources are not the only sources 01
and in animal manure were derived from national databases of nitrogen an0 phospnorus, they are believe0 to represent the
felrilizer sales for 1985-1991 (9and from Census of Agriculture majority t

d stream loads wer

rces were appor-

sheds. For example, animal manure accounted for

more than half of the nitrogen inputs in Virginia's
Shenandoah River watershed whereas, across the Po-
tomac River in Maryland, fertilizer accounted for
more than half of the nitrogen in the Monocacy River
watershed.
One interesting finding was that fertilizer nitro-
gen inputs per unit area of agricultural land were re-
lated to the dominant land use in the watershed. In
general, agricultural land in those watersheds clas-
sified as predominantly agricultural received larger
nitrogen inputs from fertilizer than agricultural land
in predominantly urban watersheds, probably as a
result of differences in the intensity of agricultural
practices. lkese findings underscore that nutrient in-
puts to watersheds are dependent on variations in
land use practices as well as the mixture of land use
in the watershed. It is not enough to know what the
land is used for. It is equally important to know the
type and intensity of that use.
Considerable uncertainty remains over the rela-
tive importance of the major nonpoint sources with
respect to their contributions to stream loads. Mass
balance studies of individual watersheds have sug-
gested that although animal manure may represent
a potentially large nutrient input to the watershed,
it may contribute only a small portion of stream load
(15-17).However, statistical studies of water qual-
ity trends have suggested that increases in total ni-
trate and phosphorus are associated with increases
in livestock population densities and fertilized acre-
age (8.18).
There is mounting evidence that atmospheric ni-
trogen inputs may be a more important nitrogen
s o w e than has been previously suggested (19).Fisher
and Oppenheimer (15) and Jaworski and his col-

VOL. 29. NO. 9.1995 I ENVIRONMENTAL SCIENCE &TECHNOLOGY rn 4 1 3 A

 leagues (16) demonstrated that nitrogen inputs from phorus transported by streams, their local impacts
the atmosphere account for 34-40% of the nitrogen may exceed those of nonpoint sources scattered
in the Chesapeake Bay and 28% in the Potomac River throughout a watershed. Consequently, any shift in
basins. Fu and Winchester (17) found similar re- focus of pollution prevention efforts to nonpoint
sults for three watersheds in Georgia and Florida. Re- sources must also be accompanied by continuing ef-
cently, Paerl (20) associated atmospheric deposi- forts to address point source impacts.
tion of nitrogen with eutrophication in coastal areas
of the North Atlantic, Baltic, Mediterranean, and Acknowledgments
North seas. This study was conducted as part of the US. Geological Sur-
Smith and others (18)found strong statistical as- vey, National Water Quality Assessment Program (NAWQA).
sociation between atmospheric deposition and to- Reviews from Charles Kratzer, Owen Bricker, Dale Robert-
tal nitrate conceneations in northern midwestern and son, Tudor Davies, and a n anonymous reviewer were very
helpful in revision of the manuscript. Most important, I want
northeastern streams, particularly in forested wa-
to recognize all the NAWQA personnel who were involved
tersheds. Stoddard's suggestion that forested water- in assembling and analyzing the data used in this study.
sheds in many areas of the eastern United States may
have already reached a saturation point with re- References
spect to atmospheric nitrogen inputs (19) could ex- National Water Quality Inventory-1992 Report to Con-
plain the significantly larger stream loads in the c:res; U.S. Environmental Protection Agency: Washing-
northeastern watersheds, where about 39% of nitro- t on, DC, 1994; 841-R-94-001.
Knopman, D. S.; Smith, R. A. Environment 1993, 35, 1&
gen inputs were attributed to atmospheric deposi- L11.
tion. Babcock, B. A,; Blackmer,A. M. j.Agric. Resour. Econ. 1992,
In some streams, particularly in or near large ur- 17, 335-47.
ban centers, point sources remain a major nutrient rrachtenberg, E.; Ogg, C. WaterResour. Bull. 1994,30, 1109-
118.
source. During the past two decades, many of the Sisterson, D. L. NAPAP Report 8, Acidic Deposition: State
point source nutrient-control efforts have concen- $Science and Technoloo, Appendix A National Acid Pre-
trated on removing phosphorus because it was be- ipitation Assessment Program: Washington, DC, 1990.
Kohout, E. J. et al. Month and State Current Emission i7md.s
lieved that the nutrient was most limiting to eu- br NO,, SO,, and VOC: Methodology and Results; Ar-
trophication. Furthermore, because of the high cost ;onne National Laboratory: Argonne, IL, 1990.
of total nitrogen removal, nitrogen controls have fo- Gianessi, L. E; Peskin, H. M. An Overview of the Enuiron-
cused primarily on converting unoxidized forms of nental Data Inventory: Methods, Sources, and Preliminary
i'esults; Resources for the Future: Washington, DC, 1984.
nitrogen to a less harmful form such as nitrate. This Smith, R. A.; Alexander, R. B.; Lanfear, K. J. In National
approach has reduced nitrogen oxygen demand and Mater Summary 1990-91-Hydrologic Events and Water
the toxic effects of ammonia, but it has not re- 7ualitY; Paulson, R. W. et al., Eds.; U.S. Geological Sur-
rey: Washington, DC, 1993;Water Supply Paper 2400.
duced the total amount of nitrogen released by many U.S. Environmental Protection Agency County-Level Fer-
point sources. In most watersheds studied, point ilizer Sales Data; U.S. Environmental Protection Agency:
sources contributed relatively small portions of nu- Washington, DC, 1990; PM-221.
National Atmospheric Deposition Program (NRSP-3)l
trients to total stream loads (Figure 5). Point sources National Trends Network NADP/NTN Coordination Of-
accounted for less than 10% of the stream loads of fice: Fort Collins, CO, 1992.
nitrogen and phosphorus in 56 and 36% of the wa- Point Source Methods Document; National Oceanic and
tersheds studied, and less than 50% of the stream Atmospheric Administration; Silver Spring, MD, 1993.
Cohn, T. A. et al. Water Resour. Res. 1989, 25, 937-42.
loads of nitrogen and phosphorus in 91 and 63% of Cohn, T. A. et al. Water Resour. Res. 1992, 28, 2353-63.
watersheds studied, respectively. This means that in Helsel, D. R.; Hirsch, R. M. Statistical Methods in Water
the majority of the streams studied, nonpoint sources Resources; Eisevier: New York, 1992.
Fisher, D. C.; Oppenheimer, M. Ambio 1991, 20, 102-
were the dominant source of nutrients. 08.
The results of this study support the argument that Jaworski, N. A. et al. Estuaries 1992, 15, 83-95.
watershed management plans need to be devel- Fu, J.; Winchester, J. W. Geochem. Cosmochem. Acta 1994,
58, 1581-90.
oped on an individual watershed basis, taking into Smith, R. A.; Alexander, R. B.; Wolman, M. G. Science 1987,
account the various hydrologic, geologic, climatic, and 235, 1607-15.
land use-related factors that influence water qual- Stoddard, J. L. In Environmental Chemistry of Lakes and
ity. Developing a more complete understanding of Reservoirs; Baker, L. A., Ed.; Advances in Chemistry Se-
ries 237; American Chemical Society: Washington, DC,
which factors have the greatest impact on water qual- 1994; pp. 223-84.
ity, and the timing of those impacts, is essential to Paerl, H. A. Can. J. Fish. Aquat. Sci. 1993, 50, 2254-69.
providing effective pollution prevention and con-
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sent only a small portion of the nitrogen and phos- cal Survey Water Resources Division i n Reston, VA.

414 A VOL. 29, NO. 9, 1995 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

 Annals of Agricultural Sciences 64 (2019) 222–236

HOSTED BY Contents lists available at ScienceDirect

Annals of Agricultural Sciences

journal homepage: www.elsevier.com/locate/aoas

Significance of biochar application to the environment and economy T

a,⁎ b b
Babalola Aisosa Oni , Olubukola Oziegbe , Obembe O. Olawole
a
Chemical Engineering Department, Covenant University, Nigeria.
b
Biotechnology Research Centre, Department of Biological Sciences, Covenant University, Ota, Nigeria

A R T I C LE I N FO A B S T R A C T

Keywords: Biochar is a carbon-rich solid formed from the organic residue by pyrolysis. The productivity of biochar relies on
Pyrolysis feedstock type and pyrolysis conditions. Studies on biochar were discussed relating its application and pro-
Feedstock duction as a source of soil remediation and bioeconomy. Pyrolysis conditions, gasification, hydrothermal car-
Bio-economy bonization were discussed in this study in obtaining biochar for remediation of soil. Biochar have made sub-
Soil remediation
stantial breakthroughs in reducing greenhouse gas emissions and global warming, reducing soil nutrient
Biochar
leaching losses, sequester atmospheric carbon into the soil, increasing agricultural productivity, reducing
bioavailability of environmental contaminants and subsequently, becoming a value-added product sustaining
bioeconomy. Bio-economy implies the exploration and exploitation of bio-resources, which involves the use of
biotechnology to create new bio-products of economic value. Biochar is a marketable bio-product, which can be
used in agriculture, industries and energy sector. Thus, biochar production can enhance soil property and
provide opportunities for additional income. This review presents the production, agronomic and economic
benefits of biochar.

1. Introduction liquids. Likewise, slow pyrolysis facilitates greater produce of biochar

(36%) than fast pyrolysis (about 17%) or gasification (12%) (Uchimiya
Biochar is a solid organic residue that is obtained from biomass et al., 2011). Slow pyrolysis is also known as conventional carboniza-
pyrolysis. Biochar, being used as soil amendment have a significant tion, it yields biochar by applying heat to the biomass, which is com-
effect on soil fertility by altering the chemical, biological and physical paratively low for a long residence time (this may take days) (Cao et al.,
characteristics of the soil Awad et al. (2018). Its impact as soil 2009). However, for fast pyrolysis, biochar is produced at higher tem-
amendment enhances soil quality and plant growth with increased crop perature with very short residence time (1 s). The main dissimilarity
yield. Although, biochar resource, production process, soil type and amidst slow and fast pyrolysis techniques are their yields. Here, the
condition as well as the type of crop to be planted could influence its later gives a higher yield of bio-oil while the former gives higher yields
effectiveness. Mostly, biochar is produced from feedstock such as of biochar. In the production of biochar, the procedure begins with
agricultural wastes, animal manures and paper products. The im- biomass drying, where the particle is further heated in order for volatile
portance of these wastes for the production of biochar is an efficient materials to be released from the solid (Rhodes et al., 2008). The vo-
way to turn waste to useful and value-added substance (Brewer et al., latile compounds formed either may be (carbon dioxide, carbon mon-
2014). Production of biochar ranges from the small scale using a oxide, methane and hydrogen) or condensable organic compounds,
cooking stove to large scale using pyrolysis system. Pyrolysis is a such as acetic acids and methanol. Cracking and polymerization reac-
thermochemical technology for transforming biomass into biochar, bio- tions in the gas phase alter the entire product spectrum (Cetin et al.,
oil and syngas between 350 and 700 °C temperature in the absence of 2005).
air (Varma et al., 2018). However, the thermochemical processes that Gasification is responsible for increased porosity and surface area of
produce biochar in its solid form are pyrolysis and gasification biochar. It converts biomass primarily into a gaseous form such as
(Lehmann et al., 2011). (syngas) leading to a limited amount of oxidizing agent at an elevated
Pyrolysis are of two types viz., slow pyrolysis and fast pyrolysis; thus temperature (≥700 °C) (Brewer et al., 2014). During the gasification
they depend on the residence time and the rate of heating. Slow pyr- process, the oxidizing agents are oxygen, water or their mixtures. In the
olysis produces more syngas while fast pyrolysis produces more oils and production of biochar, the carbonaceous solid product is one of the key

⁎
Corresponding author.
E-mail address: babalola.oni@covenantuniversity.edu.ng (B.A. Oni).

https://doi.org/10.1016/j.aoas.2019.12.006
Received 20 August 2019; Received in revised form 28 October 2019; Accepted 18 December 2019
Available online 02 January 2020
0570-1783/ 2020 Production and hosting by Elsevier B.V. on behalf of Faculty of Agriculture, Ain Shams University. This is an open access article under the CC
BY-NC-ND license (http://creativecommons.org/licenses/BY-NC-ND/4.0/).
B.A. Oni, et al. Annals of Agricultural Sciences 64 (2019) 222–236

Table 1
Summary of the reaction process in biochar production.
Processes Temperature (°C) Residence time Biochar % yield Authors ref.

Slow pyrolysis 300–700 Hours to days 37 Gai et al., (2014)

Fast pyrolysis 500–1000 <1 s 15 Gurwick et al., (2013)
Gasification 750–1000 15–25 s 10 Leng et al. (2012)

Table 2
Feedstock and pyrolysis temperature on different biochar elemental composition.
Pyrolysis temperature Feedstock Carbon % biochar property Hydrogen % Oxygen %
Biochar property Biochar property

300 Rice husk 49.08–70.16 3.65–5.02 10.98–12.04

Wood bark
Sugar beet tailing
Empty fruit bunches
Dairy manure
Pinewood
Woodchips
Organic wastes
Plant residues
Human manure
Poultry manure
400 Rice husk 75.16–82.28 2.65–4.02 5.98–7.04
Wood bark
Sugar beet tailing
Empty fruit bunches
Dairy manure
Pinewood
Woodchips
Organic wastes
Plant residues
Human manure
Poultry manure
600 Rice husk 85.47–93.14 3.06–3.68 5.98–8.04
Wood bark
Sugar beet tailing
Empty fruit bunches
Dairy manure
Pinewood
Woodchips
Organic wastes
Plant residues
Human manure
Poultry manure

Fig. 1. SEM Image of biochar Structure working in soil (Laird et al., 2010).

interest. The water evaporation and volatile components released result Composition are shown in Table 2.
in increasing the fixed carbon contents of the solid (Novak et al., 2009). Biochar generally comprises of carbon and ash, its elemental com-
The polymerization of the organic materials in vapours and gases leads positions and properties vary due to the feedstock materials and pyr-
to minor char realization and intensifies the solid yield (Uchimiya et al., olysis conditions.
2011). This has a low degree of heating and longer residence time. The Biochar skeletal structure mainly includes carbon and minerals of
extent of product that can be obtained as a result of pyrolysis of certain several pore sizes (Lu et al., 2014). Micropores in biochar are accoun-
biomass can be influenced by the condition of the process, which is table for high absorptive capacity and surface area; however, meso-
temperature and residence time (Zimmerman et al., 2011). Table 1 pores are vital for liquid-solid adsorption processes, while macropores
shows a Summary of the reaction process in Biochar Production. are essential for bulk soil structure, hydrology, aeration and movement
Feedstock and pyrolysis temperature on different biochar elemental of roots (Freddo et al., 2012). The pattern and size of biochar pores are

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B.A. Oni, et al. Annals of Agricultural Sciences 64 (2019) 222–236

Fig. 2. Effect of surface and porosity of biochar of different feedstock. (a) Sugar beet tailing (Sun et al., 2012) (b) Rice husk (Baalousha, 2009) (c) Woodchips
(Gurwick et al., 2013) (d) Empty fruit bunches(Zimmerman et al., 2011). (e) Human manure (El-Naggar et al., 2018) (f) Wood bark (Saxena et al., 2014) (g) Dairy
manure (Lehmann et al., 2011) (h) Poultry manure (Cao et al., 2009).

Fig. 3. Porous structure of (a) biochar invites (Gai et al., 2014) (b) microbial colonisation (Ok et al., 2015).

likened to the composition (feedstock materials and the temperature) However, Fig. 2a to h shows SEM images of different biochar feedstock.
used during its formation. Scanning electron microscopy (SEM) ex- Porous structure of biochar and microbial activities are shown in Fig. 3.
amines the morphology and pore size distribution of biochar different Fig. 2 shows the effect of surface and porosity of different biochar
feedstocks (Quilliam et al., 2013). Furthermore, biochar's porous feedstock. The effective performance of largely depends on the feed-
structure contains other functional groups and several aromatic com- stock and pyrolysis condition used. Fig. 3 shows SEM images of porous
pounds from biomasses of lignin-based (Zimmerman et al., 2011). Fig. 1 structure of biochar invites and microbial colonisation.
shows SEM (scanning electron microscope) of biochar structure. Chen et al. (2012) termed biochar as a carbonaceous solid, which is

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B.A. Oni, et al. Annals of Agricultural Sciences 64 (2019) 222–236

porous, formed by the thermochemical transformation of organic ma-

Lehmann et al. (2011).

Quilliam et al. (2013).

Gurwick et al. (2013)

terials in an air-depleted atmosphere that has physical and chemical

Hale et al. (2012).

Chen et al. (2012)
Sun et al. (2012).
Sohi et al. (2017)

Gai et al. (2014).

Wu et al. (2015).
properties fit for safe and long-term storage of carbon in our sur-

Gai et al. (2014)

Gai et al. (2014)

roundings. Although, biochar exists in forms of charcoal but it is pro-
duced under certain controlled conditions, which makes the carbon in it
to be more stable and converted to usable products (Quilliam et al.,

Ref.
2013; Singh et al., 2012a). The micropores in the biochar allow sorption
of dissolved organic matter and improves microorganism's activity,

Pore diameter
which speeds up remediation of organic pollutants in soils (Hameed
et al., 2017). Baalousha (2009) further explained that biochar displays
great prospective for the speed up biodegradation, retaining soil ferti-

(nm)

2.24
3.21
3.07
1.25
2.18
2.11
3.14
3.74
1.02
2.27
2.04
lity and inactivating pesticides through abiotic means. However, bio-
char amended phytoremediation is a potential technology which can be

Total pore volume cm3/g

used to remove different contaminations in soil, although there are
several other methods, which are stated in the literature.

1.1. Biochar feedstock

The properties, which makes up biochar is largely influenced by its

0.014
0.054
0.021

0.022
0.045
0.031
0.026
0.016
0.057
0.029
0.07
pyrolysis conditions and feedstock; other factors include the rate at
which heat is transferred, temperatures, and residence time (RT) (Sun
et al., 2012). Sohi et al. (2017) reported that different feedstocks re-

Surface area mg2/g

sulted in different magnitudes of surface area, pores and functional
groups in biochars, and all these variables affect sorption characteristics
of biochars. Most biochar feedstocks are: rice husk, wood bark, sugar

55.9
67.5
40.1

44.4
62.1
39.4
55.4
84.0
42.1
51.0
102.9
beet tailing, empty fruit bunches, dairy manure, pinewood, woodchips,
organic wastes, plant residues, human manure and poultry manure. The
degree and purity of biochar methods of production and feedstock has

10.1
11.9
10.7

10.5
9.2
8.8
7.6
12.9
9.4

8.1
9.6
the capacity to influence heavy metals. Biochar may contain heavy
pH

metals (HMs), which include copper, zinc, nickel, lead, chromium,

manganese, and several organic pollutants, for example per-
Ash content (%)

fluorooctanesulphonic acid (PFOS), polycyclic aromatic hydrocarbons

(PAH), perfluorooctanoic acid (PFOA), phenol, dioxins and furans
(PCDD/F) and organic acids (Al-Wabel et al., 2013). Furthermore,
23.0
27.6
11.1
24.2
14.7
18.5
22.9
13.4
15.5
26.1
22.7
biochar from animal waste does have specific pore surface area (SBET)
that is lower compared to that of biochar obtained from the plant at
%yield

same pyrolysis condition and RT as a result of ash and high inorganic

33.5
27.8
26.2
37.4
33.5
25.4
29.2
22.1
31.8
36.1
compositions in most biochar generated from animal dung (Ok et al., 23.7
2015; Singh et al., 2012b). The composition of biochar (the amount of
carbon, nitrogen, potassium, calcium, etc.) depends on the feedstock
Residence time(h)

used and the duration and temperature of pyrolysis. As an example,

biochar produced from feedstocks that have greater contents of po-
tassium (such as animal litters) often have higher potassium contents
than biochar made entirely from wood (which often have higher carbon
5
4
4
6
4
6
6
5
5
4
4

contents) (Gurwick et al., 2013). However, pyrolysis conditions greatly

affect nutrient properties contents and so biochar should be tested on a
Rate of Heating

batch-by-batch basis to determine specific properties. In cases of heavy

metals accompanied by some contaminants, perfluorochemicals and its
C/min

constituents in biochar can be influenced by the purity of the feedstock,

500
500
10

10
5

7
5
5
7
4
7

but for polycyclic aromatic hydrocarbons, its compositions could be

o
Properties of biochar prepared by various feedstock.

governed by the methods of producing biochar (Hale et al., 2012).

Limited oxygen

Limited oxygen

Table 3 shows the effect of feedstock on the properties of produced

biochar.
Atm.

1.2. Biochar as soil improvement for remediation of polluted soils

N2
N2
N2
N2

N2
N2
N2
N2
N2

Application of biochar speeds up organic compounds bioremedia-

Temp.

tion, which aims at increasing the petroleum hydrocarbon-degrading

400
500
400
600
500
400
400
500
400
500
500
C o

microbial activities in biochar- amended soils. Metalloids may not be

eliminated from the soil; however, it can be transformed from one for to
Empty fruit bunches

another, especially from high to low concentration (Wu et al., 2015).

Sugar beet tailing

Poultry manure
Human manure
Organic wastes

When remediating metalloids, certain characteristics should be con-

Plant residues
Dairy manure

sidered, for example, adsorbing metalloids in woody plants and bioe-

Wood bark

Woodchips
Pinewood
Feedstock

Rice husk

nergy crops in farmland that is polluted, and metalloids removal by

Table 3

harvesting the metalloids accumulated biomass; metalloids transfor-

mation into lower toxic products (Gurwick et al., 2013; Gai et al.,

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B.A. Oni, et al. Annals of Agricultural Sciences 64 (2019) 222–236

2014). (organic and inorganic) in the soil through source pathway-receptor

Properties that make biochar appropriate for the removal of pol- linkage.
luted soils include negative charges, a high internal surface area and Stage I, it shows a site that is contaminated due to the absence of
degradation resistance (Budzianowskiab, 2017). Usually, as organic biochar. However, in stage II, the site is free from contamination as a
material is pyrolyzed at high temperatures, the biochar surface area result of biochar presence in the mechanism. Biochar ability in re-
also increases (Chen et al., 2012; Wu et al., 2018). A source is the site of moving some selected pesticides are shown in Table 4.
the pollutant, while the receptor is the site, which causes harm. It in- Biochar from different feedstocks has strong sorption capability of
volves the pathway, which acts as the mechanism by which pollutants different pesticides and other organic pollutants. Its sorption capability
migrates from source to the receptor. Pollutant, which moves from may exceed the natural soil organic matter by a factor ranging from 9 to
source to receptor along the pathway in adequate dosage to cause da- 99. Several studies have shown a drastic decline in organic con-
mage, the pollutant is considered a contaminant and the soil is, mea- taminants in biochar-amended soils (Chen et al., 2008). The mechan-
sured a contaminated soil (Sun et al., 2012). The most available (al- isms suggested are mostly surface adsorption and partition. The sorp-
though very expensive), (cost effective and time consuming) techniques tion of organic contaminants to biochar takes place frequently as a
to get rid of the contaminated soil is to take away the source or the result of partition in low-temperature biochar and due to surface ad-
receptor (Laird et al., 2010). These suggestions are sometimes un- sorption in high-temperature biochar because partition occurs in the
realistic and costly when extensive soil pollution arises or when the uncarbonized fraction and the surface adsorption in the carbonized
pollutant site is in active use. Pollutants may perhaps move from fraction (Keiluweit et al., 2010). When the pyrolysed temperature in-
sources to receptors, however many involve the termination of a pol- creases, it also results in an increase in carbonization thereby increasing
lutant into the soil solution by many pathways (Luo et al., 2011; Singh the biochar surface area by reducing the abundance of amorphous or-
et al., 2014). Biochar interruptions source pathway-receptor linkages by ganic matter. Thus, biochar acts on the bioavailable fraction of organic
adsorbing the pollutants on its surface and decreasing the concentration contaminants as explained earlier. Effect of this can be favourable
of pollutants in the soil solution (El-Naggar et al., 2018). Remediation is thereby decreasing pesticide residues in plants Zielińska and Oleszczuk
achievable if biochar permanently removes pollutants that enter the soil (2015) but unfavourable in terms of decreasing the effectiveness of
solution, thereby completely removing the pathway to receptors. After pesticides, which results in greater application rates of these chemicals.
sorption on biochar surface, pollutants can no longer act as a risk of The degree of biochar effect on the usefulness of herbicides largely
causing damage. The exact biochar surface area is very vital, which depends on the chemistry of the herbicide molecule and its mode of
adds to the adsorption capacity for organic compounds and metal ions action and on the ageing of biochar in soils (Ma et al., 2014). A com-
Tom et al. (2015). Fig. 4 shows biochar remediation of pollutants promise between the potentially promising effect of biochar on

Fig. 4. Biochar remediation of pollutants in the soil through a source pathway-receptor linkage: Application of Biochar for Soil Remediation ref. from Tom et al.
(2015).

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B.A. Oni, et al. Annals of Agricultural Sciences 64 (2019) 222–236

Table 4
biochar ability in removing some selected pesticides/contaminant(s).
Pesticides/contaminant Feed stock Pyrolysis temp./hr Temperature of adsorption oC Removal rate % of contaminant Reference

Pyrimethanil Rice husk 650/1 h 26 ± 2 82–84 Tom et al. (2015)

Simazine Wood bark 700/1 h 26 ± 2 71–86 Laird et al. (2010).
Carbofuran Sugar beet tailing 700/1 h 24 ± 2 62–65 Mukherjee et al. (2011).
Carbofuran Empty fruit bunches 600/1 h 25 ± 2 71–80 Zimmerman et al. (2011).
Pyrimethanil Dairy manure 450/− 22–23 82–84 Brewer et al. (2014).
Chlorpyrifos Pinewood 600/1 h 26 ± 2 67–74 Varma et al. (2018).
Pentachlorophenol Woodchips 650/1 h 26 ± 2 88–91 Rhodes et al. (2008).
Atrazine/simazine Organic wastes 400/− 23–24 47–52 Gai et al. (2014).
Chlorpyrifos Plant residues 600/1 h 22–23 42–47 Uchimiya et al. (2012).
Simazine Human/poultry manure 450/− 21–22 50–56 Gai et al. (2014).

Fig. 5. Diagram showing various mechanisms in biochar soil bioremediation.

pesticide remediation and its effect on pesticide efficacy is essential, plummeting their bioavailability (Lehmann, 2007). However, essential
which depends on the contaminant remediation goals and the com- plant nutrients definitely will be immobilized as a result of this me-
pound in question. Adding biochar to soil could result to rise in the chanism. It might be favourable, in nutrient extreme conditions, and
negative charges of the soil surface by reducing zeta potential and in- detrimental in nutrient-limited soils, thereby leading to deficiency.
creasing cation exchange capacity (Awad et al., 2018; Ma et al., 2014). When nutrients and pollutants are on the surface of biochar, it main-
This promotes the electrostatic attraction that exists amidst positively tains stability between the deficiency of nutrients and the immobiliza-
charged heavy metals and soil. However, because of the occurrence of tion of contaminants on soil that has been contaminated (Purakayastha
several functional groups for example COO and OH biochar surface, et al., 2016). Precipitation of heavy metals could lead to higher pH such
biochar at this point form complexes with heavy metals, thus as lead, caladium, zinc and copper, which may lead to decreased

227
B.A. Oni, et al. Annals of Agricultural Sciences 64 (2019) 222–236

mobilization (Hale et al., 2012). Fig. 5. Shows various mechanisms in plants as well as decreasing the bioavailability of heavy metals
biochar soil bioremediation from the development of biochar produc- (Oleszczuk et al., 2013). Research shows that between 350 or 450 °C at
tion in Malaysia, proceedings of the international symposium on bio- low temperature, biochar produced becomes acidic in nature; con-
char. South Korea (Kangwon National University, 2017). versely at a high temperature of 750 °C give alkalinity. Liew and Mohd-
Redzwan (2018). If the soil envisioned for biochar use is acidic, the
2. Biochar as a soil amendment for remediation biochar produced at 750 °C or above may be effective in neutralizing
the soil and increase its fertility. Otherwise, biochar produced at a low
2.1. Biochar impact on soil physicochemical properties temperature may be appropriate for alkaline soils to correct for alka-
linity difficulties. Biochar is very effective for soil improvement by
Although Biochar has the capacity to increase soil water-holding providing plant nutrients, for example, C sequestration (Wang et al.,
capability, its hydrophobicity can significantly affect this ability. 2018). Biochar is a prospective material used to increase soil quality
Biochar with a high pH value would cause a significant rise in soil pH and minimizes harmful effects of heavy metals at the site of storage
with neutral to basic properties but only a slight increase in soil with dynamics (Abbruzzini et al., 2017).
acidic pH Devi and Saroha (2015). The outcome of biochar on the ex-
changeable cation capacity value of soil repeatedly displays correlation 2.3. Biochar reactions in soil
with the fluctuation of Ca2+ present and the rise in pH value. Acidic
soils such as peat benefited from an increase in the pH but the rise of pH Biochar is used to improve potential C sink and soil C storage, in-
in neutral soil, as those soils in a temperate climate, inhibit the growth crease nutrient in soil retention and availability of nutrient, reduction
of pH-sensitive microbes (Ameloot et al., 2013). of nutrient leachate and sustain the stability of the ecosystem of the soil,
thereby adding aromatic structure in humus soil. Evidence regarding
2.2. Biochar's effect on soil properties these aspects are found in the biochar special issue Abney and Berhe
(2018). Previous research proposes that constituents of carbon in bio-
Biochar occurrence on topsoil has a substantial outcome on the char are extremely intractable in soils; e.g., the resident time for wood
natural surroundings, depth, porosity, affected texture, structure, con- biochar is about 10–998 times lengthier than most soil organic matter
sistency and all through the process of altering the pore-size distribu- (Bai et al., 2013). More so, insufficient data as regards the RT of biochar
tion, surface area, packings, particle-size distribution and bulk density obtained from several feedstocks are significantly different (Awad et al.,
(Jośko et al., 2013). It, however, changes the features of soil, which 2018).
directly affects the growth of the plant (Saxena et al., 2014). Biochar Current research has expressed the best application rates and pro-
existence has an effect on permeability, the reaction of soil to water, cedures, properties of adding biochar to the soil carbon sequestration
swelling shrinking, its aggregation and soil-preparation workability and accumulation of nutrients over a long period (Ameloot et al., 2013).
reaction to ambient-temperature variations. It alters soil physical Furthermore, biochar intermingling with soil microbial communities
nature, triggering an increase in the total specific surface area of the and the long-term fate, stability and toxicity in soil requires further
soil, which definitely increases the aeration and structure of the soil study. Biochar, however, is chemically and biologically stable than the
(Oleszczuk et al., 2013). Biochar stimulates the workings of mycorrhiza biomass feedstock C (Hansen et al., 2016). Biochar application to soil
fungi as follows: (i) changing the soil physical/chemical structures (ii) should increase soil sorption capacity of anthropogenic organic pollu-
tortuously altering the mycorrhizae, which affects the microorganisms tant, e.g., herbicides and pesticides in a systematically different way
of the soil in the environments, (iii) intruding with plant–fungus sig- than unstructured organic matter (Ding et al., 2016). Factors such as
nalling and allelochemical detoxification (iv) providing refugia from food climate change, low soil fertility, food security are the driving
fungal grazers (Ameloot et al., 2013). forces behind new technologies been introduced in farming systems.
Biochar porosity improves the habitation of mycorrhiza fungi as Soil amendment helps in risk reduction of contaminant transfer to
well as the soil quality (Kim et al., 2015). It boosts the anion and cation water receptor organisms (Awad et al., 2018). Biochar amendment is
exchange capacities of soil, which improves soil properties thereby very useful especially in its high stability against decay, thereby
cause increase in pH and total P and N, boosting better root improve- prolonging its lifespan in the soil which enriches the soil properties, it
ment and reducing aluminium that may be present. However, biochar also helps to retain soil nutrients, soil quality by increased pH, micro-
reduces drought by increasing the moisture content of the soil, thus bial flora etc. (Lehmann et al., 2006). Other benefits of biochar for soil
reducing soil erosion and nutrient leaching (Ma et al., 2014). Conse- amendment include: decrease of CO2 production by the addition of
quently, biochar surface comprises of several chemically active groups; biochar concentration between 4 and 62% (w/w), reduces N2O pro-
they include ketones, diols, and carboxylic (etc.), which produces vast duction levels higher than 25%. Biochar can be used for pathogens
potential for the adsorption of noxious elements such as manganese control against airborne and soil-borne pathogens reported by Sun et al.
(Mn) and aluminium (Al) in acid soils, and lead (Pb) Cadmium (Cd), (2012) which further stressed that biochar obtained from citrus wood,
arsenic (As), nickel (Ni) and copper (Cu) in heavy metal-polluted soils has the capacity of controlling airborne grey mould on Lycopersicon
Wild and Jones (2009). Some particles that exist in biochar causes a esculentum. However, addition biochar to asparagus soils polluted with
significant rise in the porosity of the soil and thus promote airflow Fusarium has reduced Fusarium root rot disease. Biochar reduces the
through the landfill cover and, thus increased oxygen diffusion that bacterial wilt in tomatoes, study shows that these biochar are obtained
arises inside the landfill cover, which may lead to high degrees of mi- from municipal organic waste and it suppresses the diseases in Ralstonia
crobial degradation. Biochar application to soil has several advantages solanacearum infested soil. Researchers explore the use of biochar
it offers (Freddo et al., 2012). As a soil conditional, biochar increases amended composts for remediation of diseases caused by fungi and
the biophysical features of the soils, for example, soil-nutrient retention bacteria in soil, they found out that the disease was suppressed by
and water-holding capacity, while stimulating the growth of a plant biochar's improvement due to the presence of calcium compounds, and
(Harvey et al., 2012). some certain properties of the soil which may be physical, chemical or
Biochar has several advantages and can be used in (a) reducing the biological in nature. Biochar helps in adsorbing nutrients, minerals
tensile strength of the soil, (b) increasing pH and soil structure, (c) present in the soil and pesticides thereby preventing the uptake of these
increasing the efficiency of fertilizer use, (d) reducing the aluminium chemicals into water bodies and the degradation of these waters from
toxicity to plant roots and microbiota (e) increasing soil conditioning agricultural activities (Bai et al., 2013) (Table 5).
for earthworms inhabitants (Sun et al., 2012). Biochar further decreases Biochar may greatly diminish toxicity and transport of common
the leachate of soil nutrients, thereby improving nutrients available for pollutants in the soil through decreasing their bioavailability and

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Table 5
Organic and inorganic toxic compounds in biochar.
Biochar Pyrolysis Temp. (°C) Toxic element concentration (mg/kg) Reference (s)

Rice husk 300–500 0.05–6.2 PAHs; 0.05–4.3 Cu; 0.07–0.5 Zn; 0.11–1.66 Pb; Ma et al. (2014); Hale et al. (2012); O'Toole, et al.,
0.02–0.020 Cr (2013).
Wood bark 200–800 3.07–24.61 PAHs; ~0.03 Cd; 0.01–0.040 Cr; 0.05–0.14 Zn Devi and Saroha (2015); Awad et al. (2018).
Sugar beet tailing 100–400 52–120 PAHs; 0.14–2.17 Cd; Bai et al. (2013).
Empty fruit bunches 400–800 27–66 PAHs; 0.05–0.11 As; 0.01–0.04 Pb Ma et al. (2014); Wang, et al., (2018).
Dairy/poultry/human manure 230–650 0.4–30 PAHs; Abbruzzini, et al., (2017)
Pinewood 230–700 15–160 PAHs; 0.02–0.31 Zn Novak et al. (2010); Ahmed et al. (2016).
Woodchips 230–800 12–105 PAHs; 0.02–0.037 Cr Saxena et al. (2014); Yin et al. (2014).
Organic wastes/plant residues 230–800 3.4–86 PAHs; 0.07–0.5 As; 0.12–1.75 Pb Zeng et al. (2018); Naisse et al. (2015).

corncob biochar desorbed 175 mg/kg of PO3−4 (Hale et al., 2012;

Herath et al., 2015). Consequently, biochar desorption properties solely
depend on the pyrolysed temperature, feedstock type, and the rate
biochar application. Thus, it is believed that several biochar types
should be able to accomplish different soil nutrients in the same soil, or
can be used differently in soils to obtain the anticipated nutrient supply
effects (Lehmann and Joseph, 2015; Roy and McDonald (2013).

3. Biochar application for gas remediation

Biochar is very active for the remediation of poisonous substance

from gas. Biochar from camphor, rice hull, bamboo, sludge, hardwood
chip and manure from pig effectively eliminates H2S from biogas with
adsorption capacity between 110 and 370 mg H2S/g biochar, which has
efficiency removal of over 96% Lehmann and Joseph (2015). Adsorp-
tion of H2S was essentially assisted by biochar moisture content
(> 85% v/w), pH (> 8.0), existing surface area and chemical bonding
with surface radical groups, for example, OH and COOH. H2S inter-
mingling with alkali biochar surface through ionic interaction with
eOH and eCOOH functional groups in the presence of oxygen and
water which result to the formation of (K, Na)2SO4, and maybe bioa-
vailable as SO42− to plants (Lehmann and Joseph, 2015). Biochar ob-
tained from peanut shells (PBC) and soybean straw (SBC) at pyrolysed
temperatures between 350 and 750 °C was tested for trichloroethylene
Fig. 6. Improvement in biochar benefit to the environment. abstraction. Biochar of (PBC) and (SBC) are obtained at a high pyr-
olysed temperature (SBC700 and P-BC700) were more efficient than
increasing their localised build-up, however, the rate and degree are not those obtained at a low pyrolysed temperature (SBC300 and PBC300)
well understood. Biochar can moderate movements of some inorganic for trichloroethylene elimination, and were compared to the market-
and organic contaminants in the soil (Awad et al., 2018). The large able activated char. The rate of application of 0.28 g/L PBC700 was
surface area of biochar could be toxic to soil fauna, thus, it can act as a better compared to that of activated biochar with removal effi-
microbial habitation. Salleh (2017) shows the benefit of biochar to the ciency > 88% up to trichloroethylene concentration of 9 mg/L. The
environment from the development of biochar production in Malaysia, efficiencies elimination of biochar (PBC700 and SBC700) was credited
as concluded in Fig. 6. to enhance the hydrophobicity (~10percentage O removal) and surface
area (12–410 m2/g), and reduced the polarity of biochar in comparison
with (PBC300 and SBC300) (Lehmann et al., 2006).
2.4. Release of nutrient from biochar

Several researchers show how biochar affects nutrient availability 3.1. Biochar and greenhouse gases
positively, which makes it a great prospect as a slow-release fertilizer in
the soil. When nutrients from biochar are release (especially the ad- The key cause of changes in the climate is the increase in green-
sorbed nutrients) it is solely influenced by its desorption characteristics. house gases (GHG) and global warming, but carbon dioxide (CO2)
Some of its features may have major effects on nutrient desorption from emission contributes above 77%. The carbon (IV) oxide emission over
biochar (Kuzyakov et al., 2014). Zhang et al., 2015 revealed that the soil respiration is about 10 times higher compared to that produced
rates of desorption of ammonium from hardwood biochar rise from from the burning of fossil fuel (Nguyen et al., 2010). Furthermore, it is
about 19% to 29%, due to a decrease in the pyrolysed temperatures essential to decrease carbon dioxide contaminants from agricultural soil
range from 650 to 450 °C (Harvey et al., 2012). Considering black soil, to moderate climate change. Biochar is essentially used to increase soil
the minimum per cent of P desorbed over lower P loads (19 mg/L) rises carbon sequestration and reduces nitrous oxide (N2O) emission and
from 35% to 40% with a rise in biochar application rates ranging be- (CH4) emission (Leng and Huang, 2018). Current research has shown
tween 1 and 11%. Researchers specified that above 66% of the P ad- that biochar could possibly decrease the GHGs emissions, which are
sorbed by biochar was release at higher P loadings (105 and 250 mg/L) liable for global warming, nitrous oxide and methane, from the soil,
(Kuzyakov et al., 2009). This shows that the percentage desorption of P which have major effects on climate change. Example of this can be
may increase by enhancing biochar application rates and P loadings. shown in, biochar obtained from paper mill waste, biosolids and green
Furthermore, cacao shell biochar desorbed 1487 mg/kg of PO3−4 and waste poultry litter reduces N2O emissions from an acidic Ferro sol,

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some researchers discredit the facts. These show that several biochar The decline in cation exchange capacity at a high pyrolysed condition is
types have an effect on GHG emissions from soils in several ways. It is ascribed to biochar aromatization alongside with the disappearance of
evident that water content in the soil, type of feedstocks of biochar and functional groups on biochar. Thus, biochar that is formed at low
biochar pyrolysed temperature affects biochar potential to reduce temperature has more possibility to hold fertilizer cations for example
greenhouse gases emissions (Purakayastha et al., 2016; Major et al., ammonium, which increase their utilization efficacy (Leng et al., 2015).
2010). There are several research regarding the outcome of biochar on Examples of these can be seen when an average cation exchange ca-
soil carbon dioxide emission; but the results are not conclusive because pacity determined for biochar from various feedstock (grass oak, pine,)
of the differences in research materials and methods used. There are and pH-levels (1.6–7.6) was the highest for the lowest production
steps that encompass the reduction of GHGs emissions by using biochar, temperature, 50.9 ± 14.4 c.molC/kg for produces char at a tempera-
which is multifaceted and becoming clearer methodically. Novak et al. ture of 260 °C. Pyrolysis at 450 and 700 °C gives chars with a cation
(2010) proposed that biochar application to soil intensifies the action of exchange capacity of 15.9 ± 7.0 and 20.7 ± 16, 1 c.molC/kg. Em-
microorganisms involved in the decrease of nitrous oxide to nitrogen, phasis is that several authors used different methods of production for
providing the basic mechanisms. However, the activity of nitrous oxide their chars (Santos et al., 2012).
reducing organisms is increased because of alkalinity of biochar. Like-
wise, Wild and Jones (2009) also specified that when there is an in- 4.3. pH of biochar
crease in soil potential hydrogen, the drawbacks of biochar will defi-
nitely reduce the amend soil acidity and emissions. From studies, The biochar pH is influenced by several factors, as discussed earlier,
biochar provides great adsorption sites for nitrous oxides, nitrogen due which include the types of feedstock and the processes of production/
to the large surface area, thus reducing the release of these gases from formation; however, biochar pH is neutral to alkaline pH. Biochar has
the soil ecological unit (O'Toole et al., 2013). the capability to increase the pH of a soil that has been contaminated
especially when the pH of the biochar is higher than soil pH (Rasse
4. Properties and application of biochar et al., 2017). This occurs as a result of the creation of oxides of metal
from base cations (e.g., potassium, calcium, silicon, and magnesium)
4.1. Biochar properties during pyrolysis. Thus, biochar source materials with the greatest mi-
neral concentrations give highest biochar ash components and have the
Properties that makeup biochar are porosity and its surface area, highest pH (Schulze et al., 2016). Therefore, it is expected that they will
these affect metal sorption capacity of biochar. When pyrolyzing bio- yield the greatest increase in soil pH following application. When the
mass material, micro-pores are formed in biochar as a result of water soil pH is increased, the solubility of metal cations in the soil solution
loss in the dehydration process (Yin, et al., 2016). Biochar has different decreases which makes metals to precipitate out of the solution, mostly
pore size, which may be in micro- (< 2.00 nm), macro-pores as phosphates (Zeng et al., 2018). Biochar obtained from a source such
(> 50.00 nm) and nano- (< 0.900 nm), respectively. Its porosity and as manure materials have high mineral ash contents and can restrain
surface area changes significantly with pyrolysis temperature. considerable concentrations of metals by this mechanism (Wu et al.,
Increasing the temperature from 500 to 950 °C, the porous structure of 2016). Biochar pH-value is a very important property in agricultural
biosolids biochar increases between 0.059 and 0.1 cm3/g, the surface purpose, especially in soil amendment. It is also one of the properties, in
area also increases from 25.7 to 68.9 m2/g (Zhou et al., 2017). Biochar, which case, chars from pyrolysis differs considerably from chars pro-
however, comprises of (i) moisture (ii) ash components, (iii) fixed duced via hydrothermal carbonization. Raw biomass is naturally
carbon and (iv) labile carbon and other volatile compounds. The che- slightly acidic or mildly basic in nature, which has a certain range of
mical structure of the carbon in biochar is changed when heating to pH-values from 5.2 and 7.3 (Wu et al., 2015). Additionally, ash content,
give an aromatic structure, which is greatly unaffected to microbial which is basic in nature, is increased during the process. Increase in pH-
decomposition. Consequently, biochar with C compounds is very stable value, therefore, increases the degree of carbonization (Beesley and
for a long period for as long as 100 or 1000 years. Biochar is believed to Marmiroli, 2011).
be active for long-term C sequestration (Bruun et al., 2013).
4.4. Stability of biochar
4.2. Biochar cation exchange capacity (CEC)
Biochar is very important especially for long-term soil amendment
Biochar CEC has more capability in adsorbing cations, for example, (for example in nutrient retention, carbon sequestration, and pesticide-
NH4+, Ca2+, etc., which are necessary nutrients for plants. contaminated soil remediation), however, its long-term environmental
Consequently, biochar, which has a high level of cation exchange ca- stability is not well understood (Malghani et al., 2013). Biochar stabi-
pacity, can reduce nutrient leaching losses from soils (Bruun et al., lity depends on its pyrolyzed temperatures and feedstocks. Some re-
2008). CEC expresses the total interchangeable cations for example searchers stated that some biochar could degrade speedily in some soils;
Na+, Ca2+, NH4+, K+, Mg2+), which is capable of holding a material the type of feedstock and pyrolysis conditions of the biochar can express
(Naisse et al., 2015). CEC is the negative surface charges that attract this. Researchers proposed that enhanced pyrolysis could produce
cations that describes soil fertility. The CEC is solely influenced by its steadier biochar (Smith et al., 2010). However, biochar firmness can be
surface structure, with functional groups providing surface charges, and enhanced by increasing pyrolysis temperature. For example, the stabi-
the surface area, making the surface charges available (Rahman et al., lity of sugarcane bagasse biochar increased considerably with in-
2018). Measuring the number of exchangeable ions, the material has to creasing pyrolysis temperature between 360 °C to 560 °C (Mašek et al.,
be brought into a solution. The results obtained is influenced by the 2018). Furthermore, Mašek et al. (2018) compared the stability of
value of pH to which this solution is prepared, e.g., the use of solvents biochar gotten from chicken manure to that of sugarcane; it was found
like H2O, sodium hydroxide or hydrogen chloride. Higher values of pH that the stability of chicken manure is lower than sugarcane biochar
will lead to a high value of CEC (Nguyen et al., 2008). Biochar is a (Budai et al., 2016). In conclusion, the stability of biochar rallies on the
combination of surface area and charged surface functional groups, the extent of recalcitrant carbon substrates.
high CECs produced biochar at moderately low temperatures, at which
the surface area has considerably increased compared to the feedstock, 4.5. Biochar stability evaluation
but adequate functional groups remain in the structure to provide ne-
gative charges. Therefore, biochar produced at higher pyrolysed tem- Biochar stability up until now is not well examined. Biochar effec-
perature, for example, > 500 °C has a low CECs (Rasse et al., 2017). tiveness on the enhancement of the quality of soil is not clearly

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understood on a large scale. Biochar associated with alkalinity had been

Maestrini et al. (2014); (Hansen et al.,

(El-Naggar et al., 2018); Rhodes et al.
fully neutralized and the majority of the cations from biochar had been

Spokas, (2010); Wang et al. (2018).

Zeng et al. (2018); Spokas, 2010)

lost, as a result, it built up the microbial community (Bruun et al.,
2013). Biochar addition to soil might cause some temporary changes in
a temperate agroecosystem functioning. Thus, experiments with respect

(Lehmann et al., 2006).

to biochar-amended bioremediation of heavy metals have been done in

Lynd et al. (1999)

the short term, which poses an interrogation on the long-term effec-
tiveness. Because of the ageing processes, the ability of biochar to se-

Authors Ref.
quester heavy metal decreases with time (Bruun et al., 2012). Further

(2008).
studies on ageing are needed for better clarity. However, biochar de-

2016).
gradation may be unavoidable, as it may vary from one-feedstock and
pyrolysis conditions to another. Determination of long-term stability of

Increase metabolites accumulation in soil and reduces the degradation of pollutants in

Removes heavy metals from soil, and aromatic compounds, which may contaminate

The degradation of Oxyfluorfen is rapid in soil containing biochar than soil without
biochar has not been understood if it is the type of feedstock, pyrolysis

Biochar of this kind removes PHAs and other contaminants in the soil as well as
temperature, or soil properties.

PAHs accumulation was reduced with respect to time for organic waste/plant
replenishing the bioavailability of the soil, compared to unamend soil.
5. Removal of organic pollutants

The biochar application for the removing organic pollutants from

residues. Thus its enhances soil modification on application

soil is very essential particularly for the removal of fungicide, herbi-
cides and pesticides; including insecticides chlorpyrifos, pyrimethanil,
atrazine, carbofuran, simazine, etc.) (Spokas, 2010), antibiotics/drugs
(for example sulfamethoxazole, acetaminophen, tetracycline, sulfa-
methazine, tyrosine, ibuprofen, etc.), industrial chemicals including
polycyclic aromatic hydrocarbons (PAHs) (for example catechol,
naphthalene, polychlorinated biphenyl, pyrene, m-dinitrobenzene, p-
nitrotoluene etc.), volatile organic compounds (e.g., benzene, furan,

the soils productivity in plant.

trichloroethylene, etc.), cationic aromatic dyes (e.g. methylene blue
rhodanine, methyl-violet, etc. (Herath et al., 2015). However, various
organic compounds that exist in specific waste streams (inhibitory
Remediation effect

compounds of biomass degradation such as furfural, phenolic com-

pounds, estrogen compounds in animal manure and sewage, etc., toxic
organic compounds in landfill leachate etc. (Xia et al., 2017). The ab-

biochar.
straction mechanisms are often directed by the interfaces of these
soil.

contaminants with several characteristics of biochar. The removal of

organic contaminants is predominantly through chemisorption (elec-
trophilic interaction) and physisorption, for example, hydrophobic,
2,4 dichlorophenol; Pyrene
Polychlorinated biphenyl;

electrostatic attraction/repulsion via π-π electron donor-acceptor, pore

diffusion, and H bonding) over carboxylic, diols, alcohols functional
groups (Spokas, 2010). There are other mechanisms which include;
Contaminants

partitioning, chemical transformation (through a reductive reaction or

Phenanthere

Oxyfluorfen

electrical conductivity), and the majority of the bonded contaminants

are eventually mineralized through biodegradation (by diverse micro-
PAHs

PAHs

organisms present on the surface and in the micro-pores of biochar (Yin

et al., 2014). pH, pyrolysis temperature, type of feedstock and pollutant
ratios to biochar are factors that affect biochar interactions with organic
contaminants. As higher pyrolysis temperature gives a higher surface
300 °C, 1 h

500 °C, 1 h

300 °C, 1 h

600 °C, 1 h

500 °C, 1 h
condition

area and microporosity in the biochar, it makes it desirable for the

Pyrolysis

removal of nonpolar organic pollutants, biochar formed at a lower

Biochar removal of persistent organic pollutants in soil.

temperature does not have these properties (Awad et al., 2018). The
higher the pyrolysis temperature above 500 °C, the higher the ar-
Sandy-loam
Soil tested

omaticity, low polarity and acidity on biochar result to loss of O– and

Loamy

Loamy
Sandy

H-containing functional groups. When O-bearing functional groups

Clay

decrease, it speeds up hydrophobic interactions, however, biochar

formed at a temperature < 500 °C may contain more of O– and H-
Organic wastes/plant residues/Empty

containing functional groups, which makes them have high affinity to

polar organic compounds (IBI, 2015). Examples are the removal of
Dairy/poultry/human manure
Sugar beet tailing/Pinewood/

polar insecticide and herbicide compounds, this occurs as a result of

polar specific interactions and in the process of removing different or-
ganic contaminants (e.g., agrochemicals) from the soil by specific bio-
char and mechanisms involved for their removal. Biochar has been able
fruit bunches
Woodchips

to diminish the bioavailability of organic contaminants in soil and their

uptake by plants and microorganisms (Spokas, 2010). Table 6 shows
Wood bark

Rice Hull
Biochar

biochar removal of persistent organic pollutants in soil.

Table 6

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6. Biochar benefits marketable bio-product, which can be used in agriculture, industries

and energy sector. Thus, biochar production can help build soils and
Biochar, when applying to soil help in reducing nutrient leaching. provide opportunities for additional income.
Several researchers explained that emission of N2O usually decrease by Biochar produced from forest biomass and its uses as an agricultural
about 83% after the addition of biochar, this also applies to gaseous N soil amendment have the prospect of impacting an economy, especially
losses benefits, especially soil conditioner and organic fertilizer, thus in the areas of agriculture and forest sectors, it has other impacts on
increasing carbon sequestration, soil fertility, microbial activities, value allied industries, for example, those involve is supplying biochar
of pH, recycling of plant nutrients, water holding capacity, soil con- equipment and carbon sequesters. Some researchers demonstrated that
tamination etc. (Nguyen et al., 2014; Sedlak, 2018). However, enriched the carbon dioxide sequestration payments could play a vital role in
soil physical properties, such as water storage capacity, decrease of bulk biochar profitability. Biochar can be promoted by introducing a policy
density and increase in porosity. The varied biochar composition in- on CO2 sequestration payments. In the meantime, agriculturalists may
dicate that its surface can show hydrophobic, basic, hydrophilic, acidic find it lucrative to use biochar for cultivation of cash crops that will
characteristics etc.; these make the biochar to adsorb solutes from the yield high returns. The prospective use of biochar for improving the
soil solution that affects nutrient holding. Furthermore, through ad- fertility of the soil and increasing the yield of crops looks favourable.
sorption process, biochar can increase nutrient retention. Biochar ob- However, biochar supports the retention of the soil and water, thus
tained from wood at 450 °C adsorbed about 250–520 mg P/g (McBeath reduces irrigation cost and fertilizers and recover depleted soils.
et al., 2011; Zimmerman et al., 2011). Biochar is used as a bio-se- Biochar as an additive to soil, last long and does not need to be added
questration at atmospheric condition (Roy and McDonald, 2013). Bio- every year, when compared agricultural fertilizers, thus it cost effective.
char is effective in soil conditioning for the purpose of agriculture. Due Two main products are obtained from biomass pyrolysis, they include:
to agricultural activities, organic carbon merely deteriorates on a daily syngas and oil. These products can be used as fuel, providing clean,
basis (Xia et al., 2017). The occurrence of soil organic carbon is very renewable energy. The primary application of syngas is to be used as an
vital for the sustenance of agricultural yields, alongside with the input in the pyrolysis process, either as a heating source in the drying
holding of nutrients and water, particularly potassium, phosphorus, and process or in the kiln itself. Oil, which is a by-product of biomass, is
nitrogen, which gives a habitation for soil microbes that advance soil used as a stand-alone product; however, the emissions can be heavy in
structure. Biochar as an adsorbent can be used in removing toxic pol- particulates and black carbon. Thus, it is important to upgrade the oil to
lutants from soils that have been polluted. Organic carbon content a biodiesel for fuel transportation.
present in biochar, depending on the material source, and can be 90%
high, and this material has increased carbon sequestration application 7.1. Economic importance of biochar
activities (Ippolito et al., 2017).
Transportation is considered a major factor on the economics of
7. Biochar and bio-economy biochar production. Palma et al. (2011) explains the costs effect in
producing biochar in one place and then transported to another. They
Bio-economy implies the exploration and exploitation of bio-re- measured the economic feasibility of mobile pyrolysis facilities through
sources, which involves the use of biotechnology to create new bio- an exploration of two biochar types in three states that move at dif-
products of economic value (Bugge et al., 2016). In this case, the ferent number of times. Monte Carlo financial simulation model was
feedstock is the bioresources while the bio-product is the biochar, used with transportation logistics included in their analysis based on
which are a very important feature of bio-economy. Several bio-pro- geographic information systems (GIS) data, they resolved that the net
ducts can be produced from any versatile bioresource to add value to present value of biochar increases as the number of times the mobile
the bioresource promptly (Budzianowskiab, 2017; Leng et al., 2012). In pyrolysis facility is moved drops. Fig. 7 show the economic benefits of
essence, the production, marketing, awareness campaigns and com- biochar usage.
mercialization are imperative for the sustainability of bio-economy. All The production of biochar is gaining attention due to its promising
these activities in bio-economy is an avenue for employment opportu- potentials in energy and environment. Harsono et al. (2013) showed the
nities both direct and indirect employment. In addition, quality, safety economic assumption results for the production of biochar in Selangor
and quantity of the bio-product have a significant effect on the bio- at 532.00 US$/year, and the total revenue from biochar sale was 8012
economy (Fuentes-Saguar et al., 2017). As for biochar, effective large US$/year, from Table 7. Thus, the net present value for biochar pro-
production will result in agronomic and economic benefits. For in- duction, which was determined by the amount of investment and the
stance, the yield of the crops to which the biochar is applied and the net revenue, showed positive results, of economic feasibility of biochar.
profit made because of surplus harvest determines the economic bal- Shabangu et al. (2014) reported that the cost-effectiveness of biochar
ance. This indicates that choice of crop, type of soil, and biochar ap- production depended on its selling price, with a break-even of about
plication both in quality and in quantity are very vital as it influences $220/t for pyrolysis at 300 °C and about $280/t for pyrolysis at 450 °C.
the economic balance (Baidoo et al., 2016; Shareef and Zhao (2017). Roberts et al. (2009) demonstrated that the yard waste was confirmed
Globally the demand for energy is increasing as the world popula- to be promising biomass feedstock for the production of biochar with
tion increases, since energy is required for all sectors in any nation. net margin of $69 and $16 for the high and low revenue scenario of
Presently, fossil fuel is the major source of energy. Biomass is con- CO2e, including livestock manures, cattle and horse. Therefore, biochar
sidered to be a promising renewable energy source which can be con- production can be attractive if the proceeds of the above values offset
verted into through mechanical, biochemical, physical and thermo- the economic costs of raising, harvesting, hauling and storing the bio-
chemical processes. Thermochemical conversion gives high quality and mass feedstock, alongside those of employing pyrolysis, transportation
product yield efficiency, breaking chemical bonds of organic matters and application of biochar. From the analysis, the net margin of pro-
and converts it into biochar, bio-oil and syngas. In recent times, bio- ducing biochar could be improved with cheaper feedstock and a pro-
mass have been converted into biochar, due to its economic benefits, mising processing technology (Bugge et al., 2016).
sustainability advantages and ever-increasing demand in environmental
and energy fields. 8. Further research demand
Bio-economy has been a viable strategy to refine and upgrade
bioresources in developed countries (Bugge et al., 2016; Maestrini Interest in biochar application for removal of pollutants from the
et al., 2014) while there is need to create awareness in most developing soil for the replacement of activated carbon, which is quite expensive, is
countries and encourage bio-product production. Biochar is a fast growing. Since the quality of biochar and its performance varies in

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Fig. 7. Economic benefits of biochar. Source: (a) sugar beet tailing (Sun et al., 2012) (b) Rice husk (Baalousha, 2009) (c) woodchips (Gurwick et al., 2013) (d) empty
fruit bunches(Zimmerman et al., 2011). (e) Human manure (El-Naggar et al., 2018) (f) wood bark (Saxena et al., 2014) (g) dairy manure (Lehmann et al., 2011) (h)
poultry manure (Cao et al., 2009).

Table 7 soils, and it is very important for surface charges development and
Economic analysis of biochar production (Harsono et al., 2013). collective ability of chars, which enrich water and nutrient retention in
S/N Parameter Unit Value (Approx.)
soils. The International Biochar Initiative (IBI) standards should be
strictly adhered to in order to achieve quality, safe and sustainable
1 Investment US$ 1266:00 biochar for soil remediation. In addition, a continuous update of stan-
2 Remaining value US$ 126:00 dards is very important (Yin et al., 2014).
3 Total cost US$/year 524:00
Biochar application could increase pollutant residues in the
4 Total fixed cost US$/year 170:00
5 Total revenue US$/year 532:00 amended soils. Thus, the long-term environmental fate of the seques-
6 Total variable cost US$/year 354:00 tered pollutants is not well examined. Biochar capability to sequester
7 Net present value US$ 130:00 contaminants reduces with respect to time as a result ageing (Wu et al.,
8 Break-even point t of biochar 901:00
2018). Detail knowledge of the ageing process is paramount for future
9 Benefit/cost ratio – 1.00
10 Internal rate of return % 8.96 research. Evidence of this is important in order to determine the bio-
11 Payback period year 9.97 char application rate and its occurrence in advancing remediation ef-
12 Return on investment % 17.59 ficiency (Yin et al., 2014). Much has not been done to show biochar
ability in reducing the bioavailability and leachability of the con-
taminants in soils through sorption process; however, it may facilitate
its features as regards the feedstock type and pyrolysis condition, there speedy dissipation of some organic pollutants in soil. Additional studies
is a need for future research demand in biochar development for soil are necessary to ascertain the possibility of biochar-assisted dissipation
improvement (Li et al., 2017). Removal of the contaminant is directed of organic pollutants. The immobilized microorganism technique with
by interfaces with functional groups, which are active such as eCOOH, biochar as microbial carrier indicates the potential possibility for re-
ROH and eOH of biochar. Pollutants from organic source is eliminated moving soils polluted with organic contaminants (Lehmann et al., 2006;
by electrostatic attraction/repulsion, hydrophobic through π-π electron Lu et al., 2014). Biochar development for production of optimum car-
donor-acceptor and partitioning, while the contaminants obtained from rier is important.
the inorganic source are discarded through surface-complexation, ionic- Further research on the nutrient dynamics prediction in biochar-
interactions, ion exchange and precipitation (Smith et al., 2010). For amended soils by improving the available kinetics models both infield
Biochar to be effectively improved, its properties could be tailored into and laboratory conditions. For nutrient dynamics is important to know
the removal of specific contaminants. Major research has focused on the mechanisms that affect soil nutrient availability and fertility
sequestering carbon and the improvement of soil fertility. Further at- (Saxena et al., 2014). It is important to identify several/major activa-
tention should be focused on ways in optimizing biochar use, so as to tion processes, adsorption and desorption mechanisms for different
reduce CO2 and other greenhouse gases emissions in our environment. contaminants. Microbial communities and their distribution in biochar-
It is essential to decrease carbon dioxide contaminants from agricultural amended soil have not been well understood, most importantly with
soil to moderate climate change (Wang et al., 2018). respect to the properties of biochar, for example, ion exchange capacity,
Ippolito et al. (2017) shows that, the rate at which fertilizer is used particle size, microporosity, nutrient content, pH). With several atten-
when combined with biochar should be improved, in order to have a tion giving to soil pollution and infertility, biochar application may
reduced greenhouse gases emissions and maximum soil fertility. The bring new prospect both in remediation as a source of micro and
features of biochar vary as stated above. Future research should focus macronutrients in nutrient-deficient soils (Smith et al., 2010).
on the optimization production systems to produce designer biochar Further research on the long-term effect of biochar on soil fauna, as
products that can be effectively used specifically for remediation work some harmful effects on microorganisms and earthworms, have been
(Ahmed et al., 2016). detected in laboratory analysis. This may draw attention when using
Most researches have a keen knowledge of char-weathering rates in

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B.A. Oni, et al. Annals of Agricultural Sciences 64 (2019) 222–236

high surface area biochar, which has high affinities for organic con- cadmium and zinc by biochar. Environ. Pollut. 159, 474–480.
taminants Abney and Berhe (2018). Brewer, C.E., Chuang, V.J., Masiello, C.A., Gonnermann, H., Gao, X., Dugan, B., 2014.
New approaches to measuring biochar density and porosity. Biomass Bioenergy 66,
Evidence have shown that biochar may address some constraints of 176–185.
soil to crop production; the mechanisms to this effect are divergent, this Bruun, S., Jensen, E.S., Qensen, L.S., 2008. Microbial mineralization and assimilation of
may need further explanation (Yin et al., 2014). black carbon: dependency on degree of thermal alteration. Org. Geochem. 39,
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 www.nature.com/npjcleanwater

REVIEW ARTICLE OPEN

A comprehensive review on the chemical regeneration of

biochar adsorbent for sustainable wastewater treatment
1 1✉
Tariq Alsawy , Emanne Rashad2, Mohamed El-Qelish 3
and Ramy H. Mohammed

The chemical regeneration process has been extensively applied to reactivate biochar, supporting its reusability and leading to
significant operating cost reduction. However, no recent review discusses the effectiveness of biochar chemical regeneration. Thus,
this article comprehensively reviews the chemical regeneration of biochar contaminated with organic and inorganic pollutants.
Performance of the chemical regeneration depends on adsorption mechanism, functional groups, adsorbent pore structure, and
changes in active adsorbent sites. Secondary contamination is one of the challenges facing the sustainable adaptation of the
chemical regeneration process in the industry. The paper discusses these challenges and draws a roadmap for future research to
support sustainable wastewater treatment by biochar.
npj Clean Water (2022)5:29 ; https://doi.org/10.1038/s41545-022-00172-3
1234567890():,;

INTRODUCTION (biomasses), such as agricultural and municipal wastes, are used as

Persistent water pollution is a severe global problem that feedstock to synthesize biochar. Such wastes are collected from the
threatens human beings, other creatures, and the environment environment and turned into high-value biochar adsorbents through
in the 21st century1,2. The problem aggravates with the limited different preparation methods, followed by other modification
access to safe potable water, the rapid growth of the global techniques. The main goal of the preparation and modification
population, industrial development, and climate change. Clean techniques is to produce thermal and chemical stable biochar with
water and sanitation are among the United Nations Sustainable high surface area and pore volume17. Figure 1 illustrates different
Development Goals (SDG) to be realized by 20303. Therefore, feedstocks, preparation, and modification methods used for produ-
significant research efforts have been exerted to propose cing biochar18. The chemical composition, cation exchange capacity,
innovative and proactive treatment methods to turn wastewater and moisture content of the biomass affect the yield and the
into clean water for drinking and irrigation purposes. Table 1 elemental analysis of the resultant biochar19. Table 2 lists the
summarizes the limitations of different methods used for waste- characterization of each method.
water treatment. Traditional methods such as chlorination, One of the primary benefits of biochar is its high-cost efficiency
electrochemical treatment, and membrane filtration have been owing to the low cost of the feedstock and simplicity of preparation
broadly applied in wastewater treatment4. Nevertheless, these methods20. Reactivation (regeneration) of biochar after the adsorp-
methods were proved to be inefficient and expensive5,6. tion process to be used for several cycles reduces the operational
Advanced techniques like adsorption, ultrafiltration, chemical and running cost of the wastewater treatment process significantly.
precipitation, and coagulation were developed to address the In other words, adsorbent’s high recycling and reuse are critical for
limitations of the traditional methods7. However, most of these wastewater management’s economic efficiency21,22. Therefore,
methods (like coagulation and flocculation) generate sludge that various regeneration techniques have been proposed, developed,
should be disposed of appropriately, which, in turn, negatively and evaluated to reuse adsorbents. The techniques are influenced by
affects the environment and increases the cost of the treatment the adsorbent type (either composite or pristine), its properties, and
process. Also, these ways are energy-intensive processes, making the type of adsorbate. Figure 2 shows that these techniques could be
the cost of wastewater treatment expensive8. categorized into two main groups: desorption and decomposition
The adsorption process has received particular attention owing to process23. The decomposition regeneration depends on completely
its high efficiency, simplicity, and easy operation9. Its operational and mineralizing the adsorbed pollutants or transforming them into less
capital costs depend mainly on the adsorbent cost, which is a porous toxic byproducts24. This allows restoring the adsorption capacity of
material with a high surface area and large porosity (i.e., pore biochar. Oxidation and ultrasound decomposition methods usually
volume)10. Consequently, using low-cost and eco-friendly adsorbents use high thermal energy, so they could be considered a thermal
is the main challenge facing the broad implementation of the decomposition approach. In contrast, desorption regeneration
adsorption method in wastewater treatment processes11. Recent methods are based on breaking the bonds between biochar surface
research works focus on converting low-value waste biomass to and contaminants by thermal or non-thermal methods. Thermal
high-value adsorbents for water purification12. Biochar, sometimes regeneration is a destructive method used only if the adsorbate is a
called pyrochar, is carbon-rich material extracted from waste and waste, so there is no need for its recovery25,26. High energy
biomass materials such as wood, leaves, manure, and tree roots. It is consumption in the thermal regeneration (temperature up to 800 °C)
widely utilized in adsorbing organic and inorganic contaminants increases the process cost, which could reach up to 50% of the total
from effluent waters, such as biopharmaceuticals13, heavy metals14, cost of production of new biochar27. The non-thermal methods
pesticides15, and pentachlorophenol16. Various organic wastes preserve the adsorbents and can be reused for several cycles. Among

1
Department of Mechanical Power Engineering, Zagazig University, Zagazig, Egypt. 2Department of Environmental Sciences, Faculty of Science, Alexandria University, Alexandria,
Egypt. 3Water Pollution Research Department, National Research Centre, 33 El Buhouth St., Dokki, Cairo, Egypt. ✉email: rhamdy@zu.edu.eg

Published in partnership with King Fahd University of Petroleum & Minerals

 T. Alsawy et al.
2
these methods, the chemical one is widely applied according to presents the number of research papers presented in the Scopus
published papers analyzed in the following discussion28. database29 that discussed the regeneration process of biochar. It is
found that the total number of these papers is 117 articles (over
the last 10 years); 63 of them are focused on the chemical
METHODS regeneration process. Analyzing the published papers (about 2200
The number of publications that discuss biochar applications in papers), it is found that open literature has one review paper
wastewater purification continuously increases, as proven in Fig. 3. about the regeneration of bioadsorbents like biochar30. Another
As can be seen, the cumulative number of research and review review paper presented a concise section, without details, about
papers published till 2021 is around 2200. The number of research the regeneration process of biochar used for arsenic removal31. To
papers published in 2013 and 2021 was 26 and 675, respectively the authors’ knowledge, no review paper focuses explicitly on the
(about 26 times during the last 8 years). Because biochar received chemical regeneration process of biochar, which is critical for the
much attention, many review papers were published to discuss sustainability of wastewater treatment. It is worth mentioning that
the performance of biochar in removing contaminants from the same trend and observation are found using the Web of
effluent waters. The total number of review papers about using Science database. Therefore, this work comprehensively reviews
biochar in wastewater treatment reached about 226 articles by the chemical regeneration process that generates biochar after
2021, compared to one review paper in 2013. adsorbing organic pollutants and heavy metal ions. The paper
Biochar should be reactivated (regenerated) to be reused discusses the factors that affect process efficiency. It also
instead of burning or disposing it in the environment and demonstrates the challenges and future prospects of biochar
introducing another problem. Therefore, the regeneration process regeneration.
is critical because it determines the practicability and economic The paper is constructed into three sections, besides this
feasibility of using biochar in wastewater treatment and even section and the introduction. The main section (Section #3) of the
controls the cost of the whole treatment process. Figure 3 also present review is devoted to the chemical regeneration process as
an effective method for biochar reactivation. Factors affecting the
Table 1.Limitations and drawbacks of different technologies used in regeneration process are discussed in Section #4. Different
1234567890():,;

wastewater treatment. approaches for managing secondary pollutants are mentioned

in Section #5. The challenges and outlook of biochar as a
Wastewater treatment Limitations/disadvantages sustainable adsorbent for wastewater treatment are highlighted in
technique Section #6. Finally, the conclusion is drawn in Section #7.
Adsorption - Some adsorbents are expensive.
- Regeneration process.
CHEMICAL REGENERATION
- Adsorbent disposal.
- Time-consuming. Chemical regeneration is one of the most popular ways to
Coagulation and flocculation - Operate at low pH so, alkaline should regenerate adsorbent materials32. It relies mainly on adsorbate
be added. concentration and interaction forces on the adsorbent20. The
- Not applicable with all soluble chemical regeneration process uses solvents and chemical
contaminant. reagents to desorb the adsorbate from the biochar. Organic
- Result secondary pollutants (e.g., adsorbents with limited thermal stability and low boiling point are
sludge). more suitable for chemical regeneration33. Several solvents
Reverse osmosis - Energy-intensive. (organic/inorganic) were used to desorb both organic/inorganic
- Not practical for pharmaceuticals and adsorbates34–37. Thus, in this work, the chemical regeneration is
volatile organics.
- Treated water is acidic.
classified based on pollutant type, not on the solvent type. The
regeneration of inorganic pollutants (heavy metals) and organic
Distillation - Not efficient for contaminants having
pollutants (dyes and antibiotics) are discussed and critically
boiling point >100°C.
- Most of pollutants remain behind. analyzed. Regeneration data for each pollutant is collected from
two approaches. The first type is regeneration studies, discussing
Microfiltration/ultrafiltration/ - Not suitable for dissolved inorganics.
nanofiltration - Rapid fouling and plugging. regeneration efficiency for each cycle. The second type is
- Energy-intensive process. desorption studies that discuss the desorption efficiency of each
- Pretreatment process is needed. pollutant from the biochar at each cycle. Both types of studies are
- Regular cleaning is needed. presented and critically analyzed in this work. Regeneration
Biological processes - Limited to municipal and (removal) efficiencies are more expressive than desorption
biodegradable wastewaters. efficiencies. This is because the desorption experiments do not
- Time consuming. include the performance of the next adsorption cycle (affected by
- Challenging to be controlled. the degradation in biochar). However, regeneration (removal)
efficiencies include this effect, and they measure the adsorption

Fig. 1 Feedstock, preparation, and modification techniques of biochar. The three main steps should be followed to synthesize activated
biochar.

npj Clean Water (2022) 29 Published in partnership with King Fahd University of Petroleum & Minerals
 T. Alsawy et al.
3
Table 2. Thermochemical conversion techniques and their process conditions.

Temperature (°C) Reaction time Biochar yield (%) Liquid yield (%) Syngas yield (%) Ref.
86
Hydrothermal carbonization 180–300 1–16 hr 50–80 5–20 2–5
87
Torrefaction 200–300 10–60 min 80 0 20
Flash carbonization 300–600 <30 min 37 — — 88

89
Pyrolysis 300–1000 min-hr 12–35 30–75 13–35
90
Gasification 750–900 10–20 sec 10 5 85

aqueous solution38. The regeneration was done by washing it with

2 M NaOH for 1.5 h for 5 adsorption/desorption cycles. The
removal efficiency after the first cycle was around 83%. Progres-
sion in cycles made a slight decrease in removal efficiency in each
cycle. After the fifth cycle, the efficiency was around 76.5%. The
interaction between surface co-precipitation and NaOH may be
responsible for Pb(II) desorption without changing the LMBC’s
chemical structure. Regeneration cycles of cotton stalk-derived
biochar polluted with Pb(II) ions using 0.01 M HNO3 for 3
adsorption/desorption cycles were conducted39. After the first
regeneration cycle, the removal and desorption efficiencies were
95% and 89.4%, respectively, and decreased to 85% and 62% in
the third cycle, respectively. During the desorption phase, the
active sorption sites held by Pb(II) were liberated from the biochar
Fig. 2 Different regeneration techniques proposed for adsor- surface. Mn-Zn ferrite/biochar composite (BC) (MZF-BC50) was
bents’ reactivation (adapted from ref. 23). Common types of
decomposition and desorption techniques have been widely regenerated after being loaded with Pb(II) using 0.2 M HCl as a
applied to regenerate contaminated adsorbents. regeneration agent in 6 consecutive adsorption/desorption
cycles40. A slight drop in efficiencies was observed after the sixth
cycle (removal efficiency: 91.2%). It was concluded that, after six
cycles, as the MZF-BC50 maintained its adsorption activity. Biochar
obtained from watermelon seeds (BC) and hydrogen peroxide
treated biochar (HP-BC) were regenerated after Pb(II) adsorption
using 1.0 M HCl for 5 successive adsorption/desorption cycles41.
There was a slight reduction in the removal efficiency, which
might be attributed to partial desorption of Pb(II) onto the surface
of the adsorbents and solid loss in the solution. During the Pb(II)
adsorption onto wheat straw biochar (WBC)42, the removal
mechanisms were surface complexation, precipitation, and cation
exchange. The surface complexation removal mechanism was due
to the polar groups existing on the biochar. Moreover, Pb(II)
adsorption by precipitation was promoted by the mineral
components on the biochar. Furthermore, the ions of these
mineral components (i.e., Ca(II), Mg(II), and K(I)) enabled the
cation-exchange mechanism contribution to the adsorption.
These functional groups initiated these mechanisms, therefore,
initiated chemisorption. The complexation and precipitation
Fig. 3 Number of publications related to biochar extracted from mechanisms will highly negatively impact the regeneration
Scopus database29. The number of papers presented research process. The highest desorption efficiency among many tested
about biochar is extracted by searching “biochar” and “wastewater” desorbing agents was around 50% only. Therefore, biochar with
in title, keywords, and abstract. Word :regeneration” is added to
extract the number of papers discussing regeneration. This word is high mineral ash content may exhibit surface precipitation and
replaced by ‘chemical regeneration’ to have the number of biochar cation exchange as the dominant removal mechanisms. At the
papers having a chemical regeneration method. Overview of the same time, surface complexation may be dominant in biochar rich
published papers about biochar with emphasis on the chemical with organic functional groups (OFGs). The content of biochar (ash
regeneration process. or oxygen groups) can be controlled by the pyrolysis tempera-
ture43, thereby, controlling the removal mechanism and regenera-
tion performance. Electron-rich functional groups like aromatic
efficiency after each desorption process. Both regeneration and compounds, and C-O groups can rise π-interactions as a dominant
removal data are discussed and reported in tables. removal mechanism44.

Cu(II) and Zn(II). Tables 5 and 6 list the chemical regeneration

Heavy metals and desorption efficiency of biochar used to treat wastewater
Pb(II). Chemical regeneration process was employed to recover contaminated with Cu(II) and Zn(II). In a fixed-bed column
Pb(II) after being adsorbed by biochar. Tables 3 and 4 show the adsorption/desorption experiment, 0.2 M HCl was used to
adsorption (removal) and desorption efficiency of biochar used to regenerate Hickory wood alkaline modified biochar (HMB) that
treat wastewater contaminated with Pb(II). Magnetic biochar/ had been contaminated with heavy metals, including Cu(II)45. The
MgFe-LDH (LMBC) was regenerated after adsorbing Pb(II) from desorption process started with a significant release (desorption

Published in partnership with King Fahd University of Petroleum & Minerals npj Clean Water (2022) 29
 4
Table 3. Reusability of biochar after adsorbing Pb(II) from wastewater through chemical regeneration process.

Adsorbate Biochar Regeneration conditions Performance (removal) Ref.

pH and Temp Time (h) Test type Agent 1st cycle No. of cycles Last cycle efficiency (%)
(°C) efficiency (%)
91
Pb(II) Oxidized Eichhornia crassipes biochar NA&30 — batch 0.5 mol L−1 HCl 52 10 52
47
mBR-C NA — fixed-bed column NaOH >85 2 —

npj Clean Water (2022) 29

0.1 M, 250 mL
70
Mg–Fe-LDH-RHA 5&NA — batch 0.5 mol L−1 HNO3 100 4 94
0.5 mol/L NaOH
38
Magnetic biochar/MgFe-LDH NA&25 1.5 batch 2 mol L−1 NaOH 83 5 76.5
32
Modified MnO2 biochar-based porous NA — batch 0.3 M HNO3 + 0.03 M NaOH 21 5 20.5, and was consistent
hydrogel (MBCG)
Virgin BC hydrogel (BCG) 3 1.5
51
Coconut shell biochar modified with NA — batch 1 M NaOH 60 5 45.3
magnesium (MgBC400)
92
β–cyclodextrin rice husk-derived biochar 5.5&20 — batch HCl 88 5 76
(BCMW-β-CD)
Rice husk-derived biochar (BC) 12 2
93
Nitrogen-doped Biochar (NBC-350–0.1) NA 1 batch 0.1 M HNO3 85 5 5
0.1 M EDTA 5
0.1 M NaOH 90
T. Alsawy et al.

Pristine biochar (BC) 0.1 M HNO3 20 5 3

0.1 M EDTA 22 4
0.1 M NaOH 23 25
74
MgO modified biochar derived from Crofton 5&NA — batch 1 M NaOH 80 3 62.5
weed (MBCW600)
21
KOH-activated biochar (KOHBC) 5&25 1 batch 0.1 M HCl 100 3 47
Douglas fir biochar (DFBC) 50 20
39
Cotton stalk-derived biochar NA — batch 0.01 M HNO3 95 3 85
94
GP NA 24 batch 10% thiourea 55 5 20
GP- TH 85 84
GP- AMT 97 89
95
Capsicum annuum L. seeds biochar (CASB) NA — batch 0.01 M EDTA 51.7 5 14.6 with structure
0.01 M HCl degradation
HNO3
96
Calcium-rich biochar (CS600) NA — batch EDTA-2NaCa 25 5 10
64
L-cy/FeOOH@PHB NA&25 12 batch CH3COONa (1 mol/L) 45 5 39
40
Mn-Zn ferrite/biochar composite with 50% BC NA — batch 0.2 M HCl 99 6 91.2
content MZF-BC50
41
Watermelon seeds biochar (BC) NA — batch 1.0 M HCl 73 5 60
H2O2 treated biochar (HP-BC) 90 5 80
22
Peanut shell biochar impregnated with 5&NA — fixed-bed column 0.2 M HCl and 4 % CaCl2 98 5 88
hydrated manganese oxide (HMO-BC)
72
Activated Typha angustifolia biochar NA — batch 0.6 M HCl 63.3 6 56.7
(NABC450)
49
Hydroxyapatite-biochar (HAP-BC) NA — batch 0.2 M HCl 100 5 68

Published in partnership with King Fahd University of Petroleum & Minerals

 T. Alsawy et al.
5
Table 4. Chemical desorption efficiencies of biochar after adsorbing Pb(II) from wastewater.

Adsorbate Biochar Desorption Ref.

Test type Agent 1st cycle Last cycle
efficiency (%) efficiency (%)
93
Pb(II) Nitrogen-doped Biochar (NBC-350–0.1) batch 0.1 M NaOH 93 85 (5th cycle)
0.1 M EDTA 44 15 (5th cycle)
0.1 M HNO3 30 18 (5th cycle)
Pristine biochar (BC) 0.1 M NaOH 78 32 (5th cycle)
0.1 M EDTA 29 18 (5th cycle)
0.1 M HNO3 35 9 (5th cycle)
21
KOH-activated biochar (KOHBC) batch 0.1 M HCl 100 47 (5th cycle)
Douglas fir biochar (DFBC) 50 20 (3rd cycle)
Calcium-rich biochar (CS600) batch HNO3 90 — 96

HCl 97
EDTA 80
EDTA-2NaCa 85
NaOH 29
H2SO4 4
HBA_H2O Pb(II) system batch HNO3 81.3 — 97

HBA_ H2O Pb(II) + PAA system 79.8

HBA_ H2O Pb(II) + TRT system 95.9
HBA_ H2O Pb(II) + PAA + TRT 75.4
HBA_ H2O Pb(II) system NaOH 13.8
HBA_ H2O Pb(II) + PAA system 11.5
HBA_ H2O Pb(II) + TRT system 13.2
HBA_ H2O Pb(II) + PAA + TRT 11.1
HBA_HCl Pb(II) system HNO3 82
HBA_HCl Pb(II) + PAA system 91.6
HBA_HCl Pb(II) + TRT system 84.6
HBA_HCl Pb(II) + PAA + TRT 74.4
HBA_HCl Pb(II) system NaOH 38.8
HBA_HCl Pb(II) + PAA system 14.8
HBA_HCl Pb(II) + TRT system 15.8
HBA_HCl Pb(II) + PAA + TRT 10.5
40
Mn-Zn ferrite/biochar composite with 50% batch 0.2 M HCl 98 89 (6th cycle)
BC content MZF-BC50
WBC (2 mM Pb) batch 0.01 mol L−1 CaCl2 10 — 42

−1
1 mol L HCl 55
0.1 mol L−1 acetic 42
acid (TCLP)
WBC_PMA (2 mM Pb) 0.01 mol L−1 CaCl2 3
1 mol L−1 HCl 45
0.1 mol/L acetic 40
acid (TCLP)
WBC_PMB (2 mM Pb) 0.01 mol L−1 CaCl2 5
1 mol L−1 HCl 50
0.1 mol L−1 acetic 40
acid (TCLP)
Peanut shell biochar impregnated with fixed- 0.2 M HCl + 4% CaCl2 100 At 10 BV — 22

Hydrated manganese oxide (HMO-BC) bed column

Activated Typha angustifolia biochar batch 0.1 M HCl 68 — 72

(NABC450) 0.2 M HCl 81

0.4 M HCl 92.5
0.6 M HCl 95.4
Treated raw jujube seeds biochar (UAJS) batch 0.1 M HNO3 70 — 48

0.1 M HCl 79 68 (3rd cycle)

Published in partnership with King Fahd University of Petroleum & Minerals npj Clean Water (2022) 29
 T. Alsawy et al.
6
Table 4 continued
Adsorbate Biochar Desorption Ref.
Test type Agent 1st cycle Last cycle
efficiency (%) efficiency (%)

0.1 M H2SO4 72 —
Coaltec Energy, Inc Biochar batch 0.5 M HNO3 65 — 53

1 M HNO3 85
1 M H2SO4 20
0.5 M HCl 45
1 M HCl 51
0.1 M HNO3 65 60 (3rd cycle)
0.1 M H2SO4 13 —
0.1 M HCl 37
Hickory wood alkaline modified fixed- 0.2 M HCl 100.6 94.4 (2nd cycle) 45

biochar (HMB) bed column

22
Peanut shell biochar impregnated with batch 0.2 M HCl and 4% 100 100 (5th cycle)
hydrated manganese oxide (HMO-BC) CaCl2

rate) peak for Cu(II), followed by a gradual decrease in the surface. As a result, HAP-BC remained stable. A composite of
desorption rate throughout the regeneration phase. According to biochar/pectin/alginate hydrogel beads (BPA) was used to absorb
mass balance calculation, the acid was 85.1% efficient in Cu(II) from an aqueous solution50. The adsorption process was
regenerating the HMB. Si–Mn binary modified biochar (SMBC) mainly controlled by chemisorption. Adsorption pH was not
polluted by Cu(II) was regenerated for 5 adsorption /desorption possible to favor the physisorption using higher pH, as the free
cycles46. The removal rate in the first cycle was 98.7% and cationic form of Cu(II) occurs at pH <6, otherwise, Cu(OH) may be
degraded to around 80% in the fifth cycle. This trend was mainly formed and precipitate on the biochar surface. In turn, heavy
because, during the desorption process, EDTA-2Na interacted with metals (especially Cu) solutions are limited to acidic pH to avoid
other metal ions in SMBC, disrupting the original structure of precipitating, thus, avoiding lower regeneration performance. This
biochar to some extent. mBR-C loaded with Cu(II) was recovered agrees with the conclusions of Cd(II) and Pb(II) desorption studies
using NaOH solution (0.1 M, 250 mL)47. The adsorption capacity of from MgBC400 using NaOH (high pH)51. Even in acidic conditions,
85% was kept after two cycles of adsorption-desorption in a fixed- precipitants may form21. Therefore, it can be concluded that
bed column. The high regeneration capacity could be attributed surface precipitation is a general—unavoidable—mechanism for
to the intensive activation of NaOH on the carbon materials, which metal cations adsorption/desorption.
improved interconnected pore structures and decreased pore
deformation, hence promoting adsorption capacity. Adsorption of Cd(II) and Ni(II). Chemical regeneration has been applied to
heavy metals (Cu(II) and Zn(II)) on ABC1 biochar was physisorption regenerate biochar saturated with Cd(II) and Ni(II). MBCG and a
and chemisorption44. Physisorption of ABC1was mainly due to BCG were regenerated to release Cd(II) adsorbed on their
pore filling effects which resulted from pHpzc of 8.2 (slightly basic) surfaces32. Both biochar materials were eluted with 0.3 M HNO3
due to the FG delocalized π electrons. The pH of the adsorption solution and then regenerated with 0.03 M NaOH for five
solution was 6. Therefore, as pH < pHpzc, the biochar was adsorption/desorption cycles. The poor removal efficiency was
positively charged, and physisorption was not favored by present (under 10%). However, the MCBG’s regeneration capacity
electrostatic attraction as it repelled the positive metal cations. remains consistent since it kept 92.1% of its original (zeroth cycle)
However, chemisorption at this pH is dominant using metal efficiency. On the other hand, comparing BCG’s last cycle
cations exchange. Heavy metal ions also reacted with OFGs via regeneration efficiency to the initial efficiency, it declined
complexation and precipitation. These strong mechanisms explain dramatically (78.4%) by the fifth cycle. The mesoporous structure
the deterioration in the removal efficiency from 90% at the first of MBCG helped to prevent the loss of inner active sites over time,
cycle to 62% after 6 cycles. resulting in better regeneration.
Zn(II) was desorbed after being adsorbed by treated raw jujube Cd(II) desorption using HNO3 from OSR550 was studied52.
seeds biochar (UAJS) using 0.1 M of HNO3, HCl, and H2SO448. The OSR550 showed the maximum Cd(II) desorption abilities under
desorption efficiencies were around 90%, 93%, and 91% for HNO3, mild HNO3 concentration, where Cd(II) and different heavy metals
HCl, and H2SO4, respectively. As HCl achieved relatively higher were desorbed from biochar obtained from Coaltec Energy, Inc
desorption efficiency, it was used for 5 adsorption/desorption using different concentrations of HNO3, H2SO4, and HCl53. Higher
cycles. The desorption efficiency at the fifth cycle was around 87% acid concentrations may degrade the biochar structure, lowering
which was not much of a reduction in the efficiency, meaning it sorption and desorption efficiency owing to biosorbent loss. 0.1 M
was stable. HAP-BC was regenerated after Zn(II) adsorption using acid concentration was low enough not to damage the biochar
0.2 M HCl for 5 cycles49. The removal efficiency was 95% for the surface and achieved high desorption efficiency for most metals.
first cycle and remained constant for the second cycle. Declination Therefore, 0.1 M concentration was used to compare different
in removal efficiency started from the third cycle and continued till acids. HNO3 was the best fit for heavy metal removal in this study.
reaching 75% in the fifth cycle. The adsorption capacity decrease Therefore, 0.1 HNO3 was used for 3 adsorption/desorption cycles,
over the cycles was attributed to adsorbent waste and degrada- and the first and third cycle Cd(II) desorption efficiencies were
tion. However, this degradation was not that much. Therefore, 72% and 68%, respectively.
after five regeneration cycles, the Hydroxyapatite-biochar (HAP- For six consecutive cycles, the regeneration efficiency of ABC1
BC) surface was rough, and most of its particles remained on the polluted with Ni(II) was investigated using 0.1 M NaOH44. After the

npj Clean Water (2022) 29 Published in partnership with King Fahd University of Petroleum & Minerals
 T. Alsawy et al.
7
Ref. first cycle, the removal efficiency was reduced to around 87% from

44

28

50

46

91

69

47

49

28

91

49
92% in the initial cycle. After the first three cycles, it was further
reduced to 80%. By the end of the sixth cycle, the reduction in
removal efficiency was around 16%. The reduction in the removal
efficiency was due to the loss of active sites and pores blockage in
the ABC1. PLB was used to adsorb heavy metals (Cu(II), Zn(II), and
efficiency (%)

Ni(II)) from wastewater28. In the first two cycles, removal

1st cycle efficiency (%) No. of cycles Last cycle

efficiencies of Ni(II) were 67% and 55%, respectively28. After the

80.24

87.70
third cycle, the removal efficiency dropped drastically to 19% and

47.5
ended at 3% by the sixth cycle. The difference in the first cycle’s
11
74

41

80

75
4
–

–
removal efficiencies of different heavy metals may be attained in
the different metal radii and pKa. Biochar had a large average pore
size of 30.15 nm, its pHpzc was around 2, and the pH of the
10 adsorption solution was 5. Thus, physical adsorption was favored

10
using electrostatic attraction. However, chemisorption was domi-
6
6
5

4
2

5
6

5
nant rather than physisorption. Therefore, the type of pollutant
Performance (removal)

matters in the prediction of removal mechanisms. Even when

physisorption is favored, chemisorption dominates. Moreover, at
alkaline pH in the adsorption process, heavy metal ions can form
precipitates and surface complexes54. Thus, the first adsorption
process in alkaline conditions may be successful. However, due to
the precipitants and complexes formed, the desorbing agents
98.76

97.87

47.5
>85

could not completely regenerate the biochar. Therefore, the

87
96
81

41

98
89

95

subsequent adsorption cycle performance deteriorates due to

Reusability of biochar after adsorbing Cu(II) and Zn(II) from wastewater through the chemical regeneration process.

active site loss and pore blockage. This trend was observed in
many studies. Tables 7 and 8 present the chemical adsorption and
desorption efficiency of biochar used to treat wastewater
fixed-bed column 0.1 M NaOH, 250 mL
EDTA-2Na solutions

contaminated with Cd(II) and Ni(II).

HCl
0.5 mol/L HCl
0.1 M NaOH

−1
(0.1 mol/L)

1 M NaOH

Organic pollutants
0.2 M HCl

0.2 M HCl

0.2 M HCl
0.5 mol L
1 M HCl

1 M HCl

Dyes. Tables 9 and 10 summarize the chemical regeneration of

Agent

biochar contaminated with different types of dyes. Sulfur-treated

Tapioca peel (S@TP) biochar was regenerated after Malachite Green
(MG) and Rhodamine B (RhB) adsorption using 0.1 M NaOH for five
cycles55. For MG dye, the removal efficiencies of the first and fifth
cycles were around 82% and 57%, respectively. However, for RhB
Time (h) Test type

dye, the removal efficiencies were around 88% and 68% for the first
batch
batch
batch

batch
batch

batch
batch
batch
batch
batch

and last cycles, respectively. With each cycle, the regeneration

efficiency of the S@TP biochar reduced, which might be attributable
to the increasing clustering of the adsorbent surface. Activated
empty fruit fibers biochar (PEF) loaded with cibacron blue (CB) from
an aqueous solution was regenerated using 0.3 M HCl for 7 cycles56.
20
24

12

12

20

The removal efficiencies for the first and seventh cycles were 99.5%
–

–
–

and 91%, respectively. Protonation of carbonyl/hydroxyl group(s) at

pH and temp
Regeneration

the surface of PEF in acidic solution might explain the high

desorption efficiency reported even after 7 cycles of the adsorption-
NA&30

NA&30
4.5&25

4.5&25

desorption process. CB was repelled by protonated PEF and

6&30
(°C)

liberated from the PEF surface into a desorption solution.

NA

NA
NA

NA
Magnetic biogas residue-based biochar NA

NA

Methylene blue (MB) was desorbed from loaded biochar particles/

polysulfone fiber membrane (MMM)57. The desorption used acid,
Oxidized Eichhornia crassipes biochar

Oxidized Eichhornia crassipes biochar

Si–Mn binary modified biochar SMBC

base, salt, organic solvent, and 1 M NaCl in an organic solvent

Hydroxyapatite-biochar (HAP-BC)

Hydroxyapatite-biochar (HAP-BC)
Biochar/pectin/alginate hydrogel

Calcite-modified biochar CAL/BC

solution. A desorption rate of 20–25% was reached in a very acidic

solution, but negligible desorption happened in an alkali solution.
pineapple leaf biochar (PLB)

Pineapple leaf biochar (PLB)

Very acidic or basic environments were not stable for desorption.

The influence of ionic strength implied that MB adsorption was not
Biochar/NaOH (ABC1)

sufficiently sensitive to salt concentration. Using 1 M NaCl or KCl

solutions, the desorption rate was <20%. Ethanol outperformed the
competition, and the rate of desorption improved as the volume of
ethanol in the aqueous solution increased.
beads (BPA)

MG dye was adsorbed on ABC1 biochar by both physisorption

(mBR-C)
Adsorbate Biochar

and chemisorption44. Physisorption was mainly due to pore filling

effects. The pHpzc of ABC1 was 8.2 (slightly basic) due to the
delocalized π electrons of the functionalities. The pH of the
adsorption solution was 6, and the pKa of MG dye was around 4.2.
Table 5.

This results in the positively charged biochar surface and the

Cu(II)

Zn(II)

negatively charged MG dye. Therefore, physisorption was favored

by electrostatic attraction. Chemisorption was involved as well.

Published in partnership with King Fahd University of Petroleum & Minerals npj Clean Water (2022) 29
 T. Alsawy et al.
8
Table 6. Chemical desorption efficiencies of biochar after adsorbing Cu(II) and Zn(II) from wastewater.

Adsorbate Biochar Conditions Desorption Ref.

Test type Agent 1st cycle efficiency (%) Last cycle
efficiency (%)

Cu(II) Manganese oxide inside a pH = 5 fixed- 0.1 M HCl + 5% 92 at 5 BV 98 at 20 BV 98

biochar (MO-L-BC) bed column CaCl2

Sb(III) 1.0 M NaOH 70 at 5 BV 92 at 20 BV
Cu(II) HFO-ABC pH = 5.5 & fixed- HCl-CaCl2 95 at 20 BV — 54

HFO-ABC Cu(II)/Cd(II) co- 4.33 h bed column

polluted
2-thiouracil- modified — batch 1 M HNO3 100 Severe degradation in — 77

biochar the modified biochar.

Hickory wood alkaline Time = 2 h fixed- 0.2 M HCl 85.1 — 45

modified biochar (HMB) bed column

Hydrated ferric oxide pH = 5.5 fixed- binary HCl-CaCl2 95 At 10 BV — 54

biochar (HFO-ABC) bed column

Commercially available 22 °C & 24 h batch 1 M HNO3 95 — 65

biochar (BCS) 0.5 M HCL 88

Wheat straw biochar BCSH 22 °C & 24 h batch 3 M HNO3 100
10 M HCL 100
Coaltec Energy, Inc Biochar 6 h batch 0.5 M HNO3 81 — 53

1 M HNO3 90
1 M H2SO4 70
0.5 M HCl 70
1 M HCl 75
0.1 M HNO3 79 70 (3rd cycle)
0.1 M H2SO4 68 —
0.1 M HCl 72
Zn(II) Hickory wood alkaline 2h Fixed- 0.2 M HCl 40.2 — 45

modified biochar (HMB) bed column

Commercially available 22 °C & 24 h batch 0.5 M HNO3 88 — 65

biochar (BCS) 1 M HCL 76

Wheat straw 4 M HNO3 95
biochar (BCSH) 6 M HCL 90
Treated raw jujube seeds — batch 0.1 M HNO3 90 — 48

biochar (UAJS) 0.1 M HCl 93 87 (5th cycle)

0.1 M H2SO4 91 —
Coaltec Energy, Inc Biochar 6 h batch 0.5 M HNO3 91 — 53

1 M HNO3 95
0.5 M H2SO4 90
0.5 M HCl 82
1 M HCl 87
0.1 M HNO3 71 68 (3rd cycle)
0.1 M H2SO4 75 —
0.1 M HCl 79

However, it involved low-impact mechanisms (hydrogen bonding Four Remazol dyes (RBB, RBV5R, RBO3R, and RBBR) were
and π-π interactions). Thus, the regeneration was much better than desorbed from green marine algae biochar using 0.01 M NaOH,
that for heavy metals. The regeneration efficiency was almost 100% 0.01 M Na2CO3, 0.01 M NH4OH, HCl, CH3OH, and 0.01 M EDTA35.
(constant) for 3 cycles and decreased to 78% in the 6th cycle. This Desorption studies revealed that acidic eluants had low desorption
emphasized the role of pH, pHpzc, and pKa on physisorption, effectiveness, whereas basic eluants had high desorption efficiency.
enhancing organic pollutants’ regeneration performance. As men- Municipal sludge biochar was regenerated after MB adsorption
tioned in the heavy metal discussion, the same ABC1 and same from an aqueous solution using ethanol/acetic acid (9/1:v/v) for
solution were used for the same positively charged metal ions and three cycles58. The removal efficiencies were around 99% and 60%
MG dye. However, different removal mechanisms arise with a whole in the first and third cycles, respectively. The removal rate of the
different regeneration performance. This indicates the role of the desorbed biochar was reduced due to the preceding desorption
adsorbate itself in determining the removal mechanism, thus processes, which solubilized some sections of the biochar, altering
regeneration performance. its surface structure and resulting in the loss/blockage of adsorption

npj Clean Water (2022) 29 Published in partnership with King Fahd University of Petroleum & Minerals
 T. Alsawy et al.
9
Ref. sites. This agreed with using 5% methanol (MeOH)59. RhB was

91

32

51

74

21

22

72

44

28

21
adsorbed into cassava slag biochar (HCS)36. The pHpzc of the HCS
biochar was 5, and the pKa of the RhB was 3.5. Therefore, maximum
adsorption occurred around solution pH = 4. As the solution pH is
efficiency (%) less than pHpzc of biochar, the biochar was positively charged. And
RhB was negatively charged because the solution pH was more
than the pKa. As a result, physisorption was the dominant
Last cycle

mechanism via electrostatic interaction. Therefore, regeneration

44.5

0.13
15.3

39.3
exhibited high performance. The regeneration efficiencies were
9.3

40
13

85

15
45
97.5% and 93% in the first and fifth cycle, respectively. NaOH has a

3
–
pH higher than pHpzc and pKa. Therefore, HCS and RhB were
1st cycle efficiency (%) No. of cycles

negatively charged, resulting in RhB repulsion. It can be concluded

that the solution pH in the adsorption process should be adjusted
to control the charge of biochar and facilitate the adsorption and
10
5

3
6

6
6
6
3
desorption process.
Performance (removal)

Antibiotics. CoFe2O4-impregnated banana pseudostem (Co-

BP350) biochar was regenerated after absorbing amoxicillin
(AMX) using ethanol for 5 adsorption/desorption cycles60. The
removal efficiencies were 100% and 92% for the first and fifth
cycle, respectively. The concentration of AMX on the biochar
44.5
10.1

28.3

61.1

45.3

surface may have choked active areas, resulting in this decrease

1.6

35
16
93

88
67
25
60
in the removal efficiency with consecutive cycles. However,
after five cycles, the percentage drop did not surpass 5.7%,
0.2 M HCl + 4% CaCl2
Reusability of biochar after adsorbing Cd(II) and Ni(II) from wastewater through the chemical regeneration process.

indicating that the Co-BP350 might be a potent adsorbent.

Cefoxitin (CFX) was desorbed from sour cherry stalk waste
modified biochar (MPCWSB500) using HCl, NaOH, H3PO4,
0.3 M HNO3 +
0.5 mol/L HCl

0.03 M NaOH

0.1 M NaOH

0.1 M NaOH

KH2PO4, acetonitrile, CH3COOH, and CaCl2 at different concen-

1 M NaOH

1 M NaOH
0.1 M HCl

0.6 M HCl

trations61. Desorption efficiencies were around 1% for 0.01 M

1 M HCl
Agent

HCl, 8% for 0.001 M NaOH, 36% for 0.01 M NaOH, 28% for 0.05 M
NaOH, 18% for 0.1 NaOH, 0% for acetonitrile, 3% for 0.01 H3PO4,
4% for 0.01 M KH2PO4, 1% for 0.01 M CaCl2 and 10% CH3COOH.
fixed-bed column

The maximum CFX desorption efficiencies were achieved with

NaOH solutions of 0.01 M and 0.05 M. Authors selected 0.01
NaOH to perform a regeneration experiment. The removal
Test type

efficiency was almost constant for 5 adsorption/desorption

batch
batch

batch

batch
batch

batch
batch
batch
batch

cycles (around 98%). Interestingly, the desorption efficiency

increased to around 72%. g-MoS2 decorated biochar (g-MoS2-
BC) was regenerated after adsorbing tetracycline (TC) from
Time (h)

different water samples using 0.2 M NaOH for 5 consecutive

cycles37. Table 11 lists the reusability of biochar contaminated
20
1

2
–
–

–
–

with different antibiotics.

pH and temp
Regeneration

FACTORS AFFECTING THE CHEMICAL REGENERATION

NA&30

4.5&25
5&NA

5&NA
6&25

2&25

PROCESS
(°C)

NA

–
–

The regeneration process includes adsorption/desorption process.

So, factors affecting either adsorption or desorption process
MgO modified Crofton weed biochar (MBCW600)
Coconut shell biochar modified with magnesium

Peanut shell biochar impregnated with Hydrated

Activated Typha angustifolia biochar (NABC450)

influence the regeneration process. These factors could be

summarized as:
Modified MnO2 biochar-based porous
Oxidized Eichhornia crassipes biochar

Functional groups (FGs)

Regeneration is affected by the mutual interaction between FGs
KOH-activated biochar (KOHBC)

KOH-activated biochar (KOHBC)

and the pollutants. Same biochar with the same FGs may interact
manganese oxide (HMO-BC)

Pineapple leaf biochar (PLB)

Douglas fir biochar (DFBC)

Douglas fir biochar (DFBC)

differently with different contaminants. For heavy metal ions, the

Virgin BC hydrogel (BCG)

dominant adsorption mechanism was, in general, chemisorption.

Biochar/NaOH (ABC1)

The removal mechanisms based on these groups were complexa-

hydrogel (MBCG)

tion with OFGs, precipitations with hydroxyl and C-O groups, and
π-interactions44. These mechanisms influence the dominance of
(MgBC400)

chemisorption over physisorption. Therefore, regenerated biochar

Biochar

losses some of its active sites every cycle due to the strong
chemical bonds in chemisorption, which negatively impacts the
regeneration process. In turn, the same FGs behaved differently
during dye adsorption. For instance, ABC1 had FGs rich in
Adsorbate
Table 7.

electrons. MG dye had nitrogen-containing-FGs deficient in

Cr(VI)
Cd(II)

electrons. So, π-interactions were promoted, and the hydrogen

Ni(II)

bonds developed as well. Therefore, complexation and

Published in partnership with King Fahd University of Petroleum & Minerals npj Clean Water (2022) 29
 T. Alsawy et al.
10
Table 8. Chemical desorption efficiencies of biochar after adsorbing Cd(II) and Ni(II) from wastewater.

Adsorbate Biochar Conditions Desorption Ref.

Test type Agent Desorption efficiency (%)
1st cycle last cycle

Cd(II) HFO-ABC pH = 5.5 & fixed- HCl-CaCl2 97 at 10 BV – 54

HFO-ABC Cu(II)/Cd(II) co-polluted time=4.3 h bed column 97 at 20 BV

Hickory wood alkaline modified Time = 2 h fixed- 0.2 M HCl 85.1 – 45

biochar (HMB) bed column

KOH-activated biochar (KOHBC) – batch 0.1 M HCl 22.5 7 (6th cycle) 21

Douglas fir biochar (DFBC) 9.5 4 (3rd

cycle)
Biogas production residue biochar 22 °C & 24 h batch HNO3 (3 M) 100 – 43

(BCU400) HCl (10 M) 100

Biogas production residue biochar HNO3 (14.5 M) 90
(BCU600) HCl (1 M) 80
Wheat straw biochar (BCS600) HNO3 (4 M) 80
HCl (9.5 M) 70
Paper mill sludge biochar (PMSB) Time = 4 h batch 0.5 M NaOH + 100 – 99

0.5 M HNO3
Hydrated ferric oxide biochar (HFO-ABC) pH = 5.5 fixed- binary HCl-CaCl2 95 At 10 BV – 54

bed column
Peanut shell biochar impregnated with pH = 6 fixed- 0.2 M HCl + 4% 100 At 10 – 22

hydrated manganese oxide (HMO-BC) bed column CaCl2 BV

pH = 5 batch 100 100 (5th cycle)
Commercially available biochar (BCS) 22 °C & 24 h batch 3 M HNO3 80% – 65

9.5 M HCL 73
Wheat straw biochar (BCSH) batch 7.5 M HNO3 91
6 M HCL 88
activated Typha angustifolia biochar – batch 0.1 M HCl 71 – 72

(NABC450) 0.2 M HCl 81

0.4 M HCl 92
0.6 M HCl 94.3
Oil seed rape biochar (OSR550) 25 °C & 24 h batch 0.05 M HNO3 43.3 – 52

30
Wheat biochar (WSP550) 30.5
26.7
Miscanthus biochar (MSP700) 12.7
11.5
Coaltec Energy, Inc Biochar 6h batch 0.5 M HNO3 95 – 53

0.5 M H2SO4 95
1 M H2SO4 99
0.5 M HCl 95
0.1 M HNO3 72 68 (3rd cycle)
0.1 M H2SO4 60 –
0.1 M HCl 70
Ni(II) Hickory wood alkaline modified 2h fixed- 0.2 M HCl 46.3 – 45

biochar (HMB) bed column

43
Biogas production residue 22 °C & 24 h batch 4 M HNO3 100
biochar BCU400 1 M HCl 100 –
biogas production residue 14.5 M HNO3 100
biochar BCU600 1.5 M HCl 90
Wheat straw biochar BCS600 4 M HNO3 98
10 M HCl 100

npj Clean Water (2022) 29 Published in partnership with King Fahd University of Petroleum & Minerals
 Table 9. Reusability of biochar after adsorbing dyes from wastewater through the chemical regeneration process.

Adsorbate Biochar Regeneration Ref.

pH and temp Time(h) Test type Agent Performance (removal)
(°C)
1st cycle efficiency (%) No. of cycles Last cycle
efficiency (%)
66
Black 5 (RB5) Wood waste biochar NA – batch 0.3 M HCl 18.23 3 14.18
0.3 M NaOH 22.3 18.3
Basic Blue 12 (BB12) 0.3 M HCl 37 25.3
0.3 M NaOH 41.9 24.3
73
Reactive Yellow Ulva Porifera biochar NA – batch NaOH 99.5 3 99.3
81 (RY81)
100
Malachite Green (MG) sulfur treated Tapioca peel (S@TP) NA – batch 0.1 M NaOH 82 5 57
biochar
44
Malachite Green (MG) biochar/NaOH ABC1 NA – batch 0.1 M NaOH 100 6 –
100
Malachite Green (MG) sulfur treated chitosan-tapioca peel (S- NA – batch 0.1 M NaOH 91 5 65
CS@TB) biochar
62
Methylene blue (MB) C. glomerata biochar NA – batch 0.1 M HCl 97 5 90
G. gracilis biochar 83 80
13
Methylene blue (MB) porous biochar hydrogel PBH NA – batch 0.1 M HCl+ 0.1 NaOH 31.67 5 80
101
Methylene blue (MB) Fe2O3/TiO2 biochar 6&NA – batch water/ethanol mixture 99.8 5 66.1

Published in partnership with King Fahd University of Petroleum & Minerals

75
Methylene blue (MB) Biochar/iron oxide (biochar/FexOy) NA – batch anhydrous ethanol 84 5 71
102
Methylene blue (MB) Switchgrass-biochar (SB600) NA – batch Methanol/acetic 20 5 5
Switchgrass-biochar (SB900) (Me + Ac) 90 47
57
Methylene blue (MB) Municipal sludge biochar NA – batch Ethanol/acetic acid (9/ 99 3 60
1: v/v)
103
T. Alsawy et al.

Congo red (CR) Cotton stalks biochar (CSB) NA – batch 3.0% HCl 56.1 5 46.1
CSB/ZnONP 95.8 89.7
102
Congo red (CR) Switchgrass-biochar (SB600) NA – methanol/acetic 14 5 4
Switchgrass-biochar (SB900) (Me + Ac) 36 8
Orange G (OG) Switchgrass-biochar (SB600) NA – batch 10 4
Switchgrass-biochar (SB900) 50 10
55
Rhodamine B (RhB) Sulfur treated Tapioca peel (S@TP) NA – batch 0.1 M NaOH 88 5 68
biochar
36
Rhodamine B (RhB) Cassava slag biochar (HCS) NA – batch 0.1 M NaCl 94.5 1 –
0.1 M HCl 96
0.1 M NaOH 97.5 5 93
100
Rhodamine B (RhB) Sulfur treated chitosan-tapioca peel (S- NA – batch 0.1 M NaOH 88 5 63
CS@TB) biochar
56
Cibacron blue 3G- Activated empty fruit fibers NA – batch 0.3 M HCl 99.5 7 91
A (CB) biochar (PEF)

npj Clean Water (2022) 29

11
 T. Alsawy et al.
12
Table 10. Chemical desorption efficiencies of biochar after adsorbing dyes from wastewater.

Adsorbate Biochar Desorption Ref.

Test type Agent 1st cycle Last cycle
efficiency (%) efficiency (%)

Reactive Yellow Ulva Porifera biochar batch 0.01 M NaOH 99.5 – 73

81 (RY81) 0.01 M Na2CO3 98.9

0.01 M NH4OH 92.6
0.01 M EDTA 21.6
0.01 M HCl 3.8
0.01 M CH3OH 2.8
Cibacron blue 3G- Activated empty fruit fibers batch 0.3 M HCl 98 83 (7th cycle) 56

A (CB) biochar (PEF)

Methylene Biochar particles/polysulfone fiber batch HNO3 19.4 – 57

blue (MB) membrane (MMM) HNO3 24.8

NaOH 0
NaOH 0
1 M NaCl 18.5
1 M KCl 16.8
100 vol% methanol 49.2
80 vol% methanol 46.4
100 vol%acetonitrile 12.1
80 vol% acetonitrile 25.9
100 vol% ethanol 56.4
80 vol% ethanol 59.1
1 M NaCl in 100 vol% 75.4
ethanol
1 M NaCl in 90 vol% 92.0
ethanol
1 M NaCl in 80 vol% 77.8
ethanol
1 M NaCl in 70 vol% 72.6
ethanol
1 M NaCl in 60 vol% 70.3
ethanol
1 M NaCl in 50 vol% 64.1
ethanol
1 M NaCl in 40 vol% 54.4
ethanol
1 M NaCl in 30 vol% 46.6
ethanol
1 M NaCl in 20 vol% 41.3
ethanol
1 M NaCl in 20 vol% 29.3
ethanol
1 M NaCl in 0 vol% 17.3
ethanol
Methylene Switchgrass-biochar (SB600) batch MeOH 7.5 – 102

blue (MB) methanol/acetic 14

(Me + Ac)
Switchgrass-biochar (SB900) batch MeOH 12
methanol/acetic 23
(Me + Ac)
Orange G (OG) Switchgrass-biochar (SB600) batch MeOH 3
methanol/acetic 2
(Me + Ac)
Switchgrass-biochar (SB900) batch MeOH 12
methanol/acetic 2
(Me + Ac)
Congo Red (CR) Switchgrass-biochar (SB600) batch MeOH 2.6

npj Clean Water (2022) 29 Published in partnership with King Fahd University of Petroleum & Minerals
 T. Alsawy et al.
13
Table 10 continued
Adsorbate Biochar Desorption Ref.
Test type Agent 1st cycle Last cycle
efficiency (%) efficiency (%)

methanol/acetic 1
(Me + Ac)
Switchgrass-biochar (SB900) batch MeOH 2%
methanol/acetic 1
(Me + Ac)
Remazol dye (RBB) Green marine algae biochar batch 0.01 M NaOH 98 – 35

0.01 M Na2CO3 98
0.01 M NH4OH 89
0.01 M HCl 5
0.01 M CH3OH 3
0.01 M EDTA 20
Remazol dye Green marine algae biochar batch 0.01 M NaOH 99 –
(RBV5R) 0.01 M Na2CO3 99
0.01 M NH4OH 90
0.01 M HCl 4
0.01 M CH3OH 3
0.01 M EDTA 19
Remazol dye Green marine algae biochar batch 0.01 M NaOH 99.5 –
(RBO3R) 0.01 M Na2CO3 99.5
0.01 M NH4OH 91
0.01 M HCl 3
0.01 M CH3OH 3
0.01 M EDTA 21
Remazol dye (RBBR) Green marine algae biochar batch 0.01 M NaOH 99.5 –
0.01 M Na2CO3 99.5
0.01 M NH4OH 90
0.01 M HCl 2
0.01 M CH3OH 3
0.01 M EDTA 18
76
Methyl violet Chitin derived biochar batch 0.5 M CH3COOH 100 100 (6th cycle)
dye (MV) 0.5 M CH3COOH 100 27 (10th cycle)

precipitation mechanisms were not dominating the adsorption, Removal mechanisms

which may be linked to the high regeneration efficiency after the The strength of removal mechanisms is generally classified as
first cycle (100%)44. Comparing the regeneration performance of chemisorption (strong) and physisorption (weak)43,65,66. Chemisorp-
the MG dye and the metal ions on the same biochar, the effect of tion precipitation and complexation have high impact mechanisms,
nitrogen-containing-FGs on the MG dye adsorption and regenera- while π- interactions and ion exchange have low impact mechan-
tion is significant. isms42,67. According to the literature, biochar adsorbs pollutants by
FGs affect the regeneration of biochar in two ways: (1) They chemisorption has lower regeneration performance than by
determine the biochar surface’s acidity and basic behavior physisorption56,63. This is mainly because pollutants chemically
(biochar charge)62. Thus, they determine the amount of physi- react with the active adsorbent sites forming complexes or
sorption by electrostatic attraction and hydrogen bonds, hence precipitate on the adsorbent surface which may not be reversible
affecting the difficulty of desorption63. The surface acidity of in the desorption process65. Thus, the adsorbent loses its active sites
biochar originates from the OFGs like carboxyl and phenyl44. While gradually each cycle because its pores are getting blocked21,45. This
basic behavior may result from either: firstly, basic FGs like may lead the biochar to lose its functionalities and mass and
nitrogen-containing-FGs (Amino) or OFGs (carbonyl). Secondly, degrade with the adsorption/desorption cycles68.
from π-electrons in aromatic rings behaving as a Lewis base. (2) The rich OFGs may drive surface complexation as the primary
Determine the type of removal mechanism (chemisorption or mechanism. Therefore, precipitations were tougher to desorb than
physisorption)54. OFGs may govern some chemisorption removal the surface complexation mechanism43. Regeneration of biochar
mechanisms, namely: (I) complexation, (II) precipitation, and (III) polluted with organic adsorbates seems more straightforward
cation exchange. Especially in metals, OFGs negatively influence than heavy metals because surface complexation mechanisms are
the desorption and regeneration of the pollutant in terms of involved, as shown in Fig. 4a. The ranges of first and last cycle
efficiency and consistency46,51,64. However, how FGs behave removal efficiencies are shown in Fig. 4b. It is found that the
depends on solution pH, biochar pHpzc, and pollutant pKa, as efficiencies of the last cycle have a wide range, indicating the
presented in Table 12. instability of the regeneration of heavy metals.

Published in partnership with King Fahd University of Petroleum & Minerals npj Clean Water (2022) 29
 T. Alsawy et al.
14
Table 11. Reusability of biochar after adsorbing antibiotics from wastewater through the chemical regeneration process.

Adsorbate Biochar Regeneration Ref.

pH and Time (h) Agent Test type Performance
temp (°C)
Cycles Removal
efficiency (%)
1st cycle Last cycle

AMX CoFe2O4-Impregnated Banana — — ethanol batch 5 100 92 60

Biochar
TC Magnetic Sludge Biochar (Fe/ — — ethanol batch 5 85 70 63

Zn-SBC) — — NaOH batch 5 93 75

TC Tap Water g-MoS2 Biochar (g-MoS2-BC) — — 0.2 M NaOH batch 5 70 64 37

TC Deionized Water — — 0.2 M NaOH batch 5 70 64

TCRiver Water — — 0.2 M NaOH batch 5 72.8 65
SDZ CSB-6–2.5 Powder Biochar — — CH3OH and CH3COOH batch 6 96 93.8 104

(mass ratio 10:1)

CFX Cherry Stalk Waste — — 0.01 NaOH batch 5 98 98 61

Biochar(MPCWSB500)
Bisphenol BCMW-β-CD 5.5 & 20 — Ethanol batch 5 48 44 92

Rice Husk-derived biochar (BC) 5.5 & 20 — Ethanol batch 5 50 43

For the organic pollutants, analysis of the reviewed studies functional groups, and their net behavior is different in each
showed that the physisorption by electrostatic attraction and solution pH. Thus, solution pH for inorganic adsorbate is crucial for
interactions (i.e., hydrogen bonds and π interactions) were both adsorption and desorption efficiencies.
dominant among organic adsorbates, giving them better and
more stable regeneration performance than inorganic adsor- Solution pH. When the adsorbate is organic, physisorption
bates55,62. Chemisorption (either π interaction or cation exchange) dominates. Thus, the solution pH should promote opposite
was also involved in dyes adsorption, as shown in Fig. 5a. charges on both biochar and adsorbate60. However, when
However, it was relatively reversible44. The regeneration perfor- chemisorption dominates in heavy metals, the pH of the solution
mance was relatively stable compared to inorganics experiencing may be acidic, promoting electrostatic repulsion. The adsorption
precipitations and complexations, as shown in Fig. 5b. The first becomes pH-independent as the chemisorption is the domi-
and last cycles have higher median removal efficiency compared nant49,50. For desorption in alkaline conditions, heavy metal ions
to corresponding values of heavy metals. Moreover, their scatter form precipitants on the biochar surface54, reducing regeneration
range is less, indicating relatively high efficiency in the last performance. Therefore, the highest desorption for heavy metal
regeneration cycle. ions may occur at acidic pH41,70,72. Figure 6 summarizes—based
on literature—the effect of solution pH, adsorbate pKa, and
Adsorbent and solution pH biochar pHpzc on adsorption/desorption performance. Figure 7
summarizes the dominant removal mechanisms found in litera-
Biochar PZC. The acidic/basic activity of biochar is expressed
ture, their corresponding driving FGs, and their impact on
using its point of zero charge (pHpzc)28. If the solution pH > pHpzc),
regeneration. It is seen that the chemisorption process is the
the biochar deprotonates (negatively charged) and vice versa69.
dominant process in inorganic pollutants removal57,73.
This affects the interaction between biochar and pollutant35,37.

Adsorbate type. it determines if its adsorption involves chemical Desorbing agents

bonding (chemisorption) or attraction (physisorption). In literature, it Based on the previous discussion, we may conclude that the metal
is found that most of the inorganic pollutants (heavy metals) desorption is as complex as its adsorption. There is no
chemically react with biochar surface50,70. However, most organic straightforward way to tell which regeneration agent is the best
pollutants are physisorbed by attractions and interactions to for each pollutant. This is mainly because the desorption process
biochar’s surface, affecting the regeneration performance62. depends on the biochar composition, FGs, its physiochemical
properties, and adsorbate43,56. However, some guidelines may be
Adsorbate pKa. Adsorbates have positive or negative charges suggested:
based on their pKa and the solution pH. When the solution
pH<adsorbate pKa, the adsorbate is positively charged. On the For heavy metals. (1) Both acidic and alkaline desorbing agents
other hand, the adsorbate is negatively charged if solution were dominant in the previous studies57,74. (2) Acidic out-
pH>adsorbate pKa37,59. Thus, we can conclude that what performed alkaline desorbing agents due to the minimal surface
determines whether adsorbate is attracted to the biochar precipitants (pore blockage), resulting from dissolving the
(adsorbed) or repelled (desorbed) is the solution’s pH and pHpzc adsorbate complexes on the biochar. (3) HNO3 was concluded
of biochar and pKa of the adsorbate. Adsorbate’s pKa may have to be better than HCl in many studies because HCl may form
one, two, or three pKa values. One pKa value is common among more precipitants and complexations, leading to metals re-
dyes59, and two and three were found in antibiotics61,63,71. In adsorption on the biochar after desorption. (4) Acidic desorbing
double pKa adsorbates, we get them positively charged, agents react with the mineral structure of biochar, thus destroy-
negatively charged, and zwitterionic. The complexity of pKa ing it every cycle. This is not noticed in the desorption efficiency.
values is mainly because organic pollutants have multiple However, there is a noticeable deterioration in the adsorption

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 T. Alsawy et al.
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Table 12. Functional groups and their corresponding removal mechanism and regeneration impact 28,44,50,53
.

Removal Regeneration
Functional group
mechanism impact
Carbonyl

High

Hydroxyl (–OH)

High

Carboxyl (–COOH)
Oxygen Functional
High
Groups

Alcohol (R–OH)
High

Carboxyl (–COOH)

Moderate

Ca(II), Mg(II), K(I),

Minerals Moderate
Na(I)

Carbon Rich Aromatics Low

Phenols

High

Amine group

Nitrogen Groups High

Published in partnership with King Fahd University of Petroleum & Minerals npj Clean Water (2022) 29
 T. Alsawy et al.
16

Fig. 4 Summary of heavy metal adsorption’s dominant removal

mechanisms and regeneration efficiencies. a The surface com- Fig. 5 Summary of organic adsorption’s dominant removal
plexation is the dominant adsorption mechanism, followed by ion mechanisms and regeneration efficiencies. a The electrostatic
exchange. b The last cycle efficiency regeneration has a wide range, attraction is the highly dominant adsorption mechanism. b The
indicating the instability of the regeneration of heavy metals. Note regeneration performance of dyes was relatively stable. The scatter
that the vertical lines present the deviation from the median value. range indicates relatively high efficiency in the last regeneration
cycle. Note that the vertical lines present the deviation from the
median value.

efficiency of the upcoming cycles. (5) Salts may not have the
highest desorption capabilities, but they will not degrade the Surface area and morphology of biochar
biochar structure. (6) Alkaline agents may have better regenera- Adsorption process via physisorption enhances using biochar with
tion performance with biochar having fewer OFGs and more higher surface area. However, no correlation was found between
carbon/nitrogen FGs38. the desorption performance and adsorbent surface area39,40,56.
High regeneration performance occurred when relatively high
For organic pollutants. (1) Alkaline dominates acidic desorbing bonding strength and low surface area existed. For the strongly
agents in terms of performance because the adsorption is mostly chemisorbed heavy metals on biochar, the morphology of biochar
physisorption (electrostatic attraction), and the desorption is after each regeneration cycle was affected because the biochar’s
electrostatic repulsion35,60,75. In most cases, electrostatic repulsions pores were blocked with precipitants, and its active sites were lost
occur at alkaline pH61,73. The data extracted from the literature (i.e., destroying its original structure)46. This biochar degradation
highlights that acidic agents, besides their relative rarity, have much may occur abruptly and entirely after the first cycle21,77 or
more scatter and instability than alkaline agents. (2) No surface gradually after that50,51. The deterioration of the biochar structure
degradation occurred in most of the studies. (3) Regeneration of may be reduced by modifying the biochar to have stronger C-C
organic pollutants is more stable than the chemisorbed metals bonds77. After the regeneration process, X-ray photoelectron
(inorganic) pollutants. (4) Acids may be used based on the pH that spectroscopy and FTIR analysis could be used to characterize
satisfies the electric repulsion when adsorbates76. biochar78–80.

npj Clean Water (2022) 29 Published in partnership with King Fahd University of Petroleum & Minerals
 T. Alsawy et al.
17
Acidic desorbing agents were mainly the cause of biochar
structure deterioration during dissolving the chemisorbed pollu-
tant complexes53. HNO3 was found to be the most involved acid in
many structural deteriorations due to its strong acidity and H+
ions interactions with the mineral composition of the biochar68.
On the other hand, HCl was reported to be benign on the biochar
surface with less impact on the morphology of the regenerated
biochar48,49. Not many morphology deteriorations were reported
for organic pollutants due to the weak bond between adsorbent
and pollutant58. Interestingly, acid erosion of the biochar after
regeneration may be beneficial because the surface area got
enlarged. This approach enhances the adsorption (physisorption)
process in subsequent cycles13.

MANAGEMENT AND DISPOSAL OF SPENT ADSORBENTS

The major limitation of chemical regeneration is secondary
pollution. Therefore, sustainable management of spent adsorbents
should be considered to rescue the environment. Different
disposal techniques were proposed, such as reuse, incineration,
and landfilling81,82. In case of reuse, the spent adsorbents are used
in applications such as soil amendment, energy conversion,
storage devices, capacitors, and catalyst/catalyst support. Studies
highlighted that nutrient-enriched biochar is an environmentally-
friendly fertilizer that could be a substitution for synthetic
fertilizer83. Metal-impregnated biochar might replace carbon
nanotubes materials, and they might be used for tar removal or
as supercapacitors soon82. Incineration is a final management
technique and is commonly referred to waste-to-energy process,
in which spent adsorbent is utilized as a source of thermal energy,
instead of coal, with reduced corrosiveness and toxic gas
emission84. Landfilling is another management method for
dumping spent adsorbents in landfills. The concentration of
adsorbate in spent sorbent should be determined before disposal
to decide whether the process is permitted. Heavy metals-
contaminated biochar requires a pretreatment process before
final disposal in a landfill. Besides these methods, other ways
could be used, such as microwave irradiation, phytocapping, and
phytoremediation85.

REMARKS, CHALLENGES, AND FUTURE PERSPECTIVES

Chemical regeneration of biochar has been recognized as a
fabulous sustainable technique that can reduce the process’s cost
and energy consumption. Chemical regeneration, unlike thermal
regeneration, is an effective regeneration method that does not
require high temperatures. Most chemical regeneration tempera-
tures range between 20 and 35°C, significantly reducing cost and
becoming a suitable method for more biochar types. There is no
straightforward rule regarding the selection of chemical regen-
eration agents. Both organic and inorganic agents can be used for
organic and inorganic biochars and still have high regeneration
performance. Literature showed that the performance of chemi-
cal regeneration depends on the reaction between pollutants,
biochar material, and chemical regeneration agents. Adsorption
mechanism, FGs at biochar surface after adsorption, adsorbent
pore structure, and changes in active adsorbent sites affect the
chemical regeneration performance. Regeneration of biochar
polluted with organic adsorbates seems more straightforward
than heavy metals. This is attributed to fewer involved mechan-
isms with physisorption dominating the adsorption. The dis-
cussed studies prove that cation exchange is a biochar common
Fig. 6 Mutual effect of pH, pKa, and pHpzc on adsorption and heavy metal removal mechanism. Most results indicated a decline
desorption. Effect of solution pH, adsorbate pKa, and biochar pHpzc in regeneration performance when the number of cycles
on adsorption/desorption performance. increases. However, when the operating parameters are opti-
mized, the decline in performance becomes insignificant.
Precipitations were shown to be tougher to desorb than surface
complexation mechanism.

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 T. Alsawy et al.
18

Fig. 7 Percentage of the dominant removal mechanisms, their functional groups, and their impact on the regeneration process. The
dominant removal mechanisms found in literature, their corresponding driving FGs, and their impact on regeneration.

Both cyclic desorption efficiencies and adsorption efficiencies modification techniques that could support the complete
are discussed. High desorption efficiency does not always mean chemical regeneration of biochar.
high regeneration performance in many cases. High desorption ● In-situ functionalization and modification of biochar resulting
efficiency means the desorbing agent may desorb most from the choice of biomass feedstocks can reflect the merits
pollutants from the biochar surface. However, it does not tell of biochar in price and source.
any information about the status of adsorbent material. The agent ● Many previous works highlighted the effectiveness of using
may degrade the adsorbent (biochar) surface, losing its active biochar in extracting and recovering valuable metals from
sites and blocking its pores. Therefore, high desorption efficiency wastewater. However, these studies were conducted at the
does not talk much about regeneration performance. In other lab-scale level. Thus, more research should be exerted to
words, it does not tell how the biochar will perform in pollutant increase the value of using biochar in wastewater treatment
adsorption with cycle progression. Despite the chemical regen- and raise the technological readiness level of such an
eration advantages, it has some drawbacks, such as the secondary adsorption/desorption method.
toxic materials that could adversely affect the biochar stability in ● Chemical regeneration depends on the oxidative destruction
removing the contaminants. of organic pollutants. However, the possible generation of
In future research works, numerous challenges need to be unidentified toxic by-products still needs more research.
considered to fully adopt this technique for practical applications:
● Most studies used biochar to adsorb and desorb a single
pollutant in synthetic wastewaters. This approach is neither DATA AVAILABILITY
realistic nor practical. Therefore, future works should consider The datasets generated during and/or analysed during the current study are available
from the corresponding author on reasonable request.
real effluent wastewater with a mixture of pollutants to
illustrate the adsorbate effects on the performance of the
regeneration process. Received: 3 February 2022; Accepted: 10 June 2022;
● Research efforts should develop low-cost technologies to
recover contaminants from spent adsorbents.
● Investigating biochar characteristics (i.e., surface morphology,
porosity, functionalities) after each regeneration cycle should
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DOI: 10.1039/D2RA06481B (Paper) RSC Adv. , 2023, 13, 220-227

Enhanced adsorption capacity of activated carbon over thermal oxidation

treatment for methylene blue removal: kinetics, equilibrium, thermodynamic,

and reusability studies†

Irwan Kurnia *ab, Surachai Karnjanakom c, Irkham Irkham ab, Haryono Haryono
a, Yohanes Andre Situmorang d, Antonius Indarto de, Atiek Rostika Noviyanti
a, Yeni Wahyuni Hartati a and Guoqing Guan f
a Department of Chemistry, Faculty of Mathematics and Natural Sciences, Universitas

Padjadjaran, Jl. Raya Bandung – Sumedang KM. 21 Jatinangor, Sumedang 45363, Indonesia.
E-mail: Irwan.kurnia@unpad.ac.id ; Tel: +62-22-7794391
b Study Center of Natural Resources, Energy and Environmental Engineering, Universitas

Padjadjaran, Jl. Raya Bandung – Sumedang KM. 21 Jatinangor, Sumedang 45363, Indonesia
c Department of Chemistry, Faculty of Science, Rangsit University, Pathumthani 1200,

Thailand
d Department of Bioenergy Engineering and Chemurgy, Institut Teknologi Bandung, Jl. Let.

Jen. Purn. Dr. (HC). Mashudi No. 1, Sumedang 45363, Indonesia

e Department of Chemical Engineering, Institut Teknologi Bandung, Jl. Ganesha 10, Bandung

40132, Indonesia
f Institute of Regional Innovation, Hirosaki University, 3-Bunkyocho, Hirosaki 036-8561 ,
:
Japan

Received 14th October 2022 , Accepted 7th December 2022

First published on 21st December 2022

Abstract
Activated carbon (AC) is an effective and inexpensive adsorbent material for dye
removal, but it cannot always be used repeatedly. Furthermore, the adsorbed dyes
with toxicity usually remain on its surface. In this study, a thermal air oxidation
process was used to modify the surface of AC and decompose adsorbed methylene
blue (MB). The behavior of this process on spent AC was investigated using TGA-DTA,
while the degradation of MB before and a"er the regeneration process was analyzed
using a carbon, hydrogen, nitrogen, sulfur (CHNS) analyzer. It was discovered that
thermal air oxidation could promote the formation of oxygenated functional groups
on AC produced from steam-activated carbon coconut shell (SACCS), which when
treated at 350 °C (denoted as SACCS-350), demonstrated an adsorption capacity 2.8
times higher than the non-air-oxidized AC (SACCS). The key parameters for the MB
adsorption of SACCS and SACCS-350, such as kinetics, equilibrium, and
thermodynamics, were compared. Moreover, the SACCS-350 could be reused at least 3
times for the adsorption of MB. Based on these results, thermal air oxidation
treatment could successfully improve the adsorption performance of AC and
regenerate spent AC through a reasonable and environmentally friendly process
compared to other regeneration methods.

Introduction

Wastes from the textile industry have led to severe environmental problems worldwide as
dyes with high intensity are difficult to decompose through only natural biodegradation
:
processes. 1 It has been reported that about 10–15% of synthetic dyes are released into the
environment through textile wastes 2 and have a serious potential to disrupt the lives of
aquatic organisms. 3 However, the most widely used cationic synthetic dyes, such as
methylene blue (MB), can also negatively impact humans and are more toxic than anionic
synthetic dyes. 4 Therefore, there is an urgent need to develop efficient methods for
removing these wastes to ensure a sustainable ecology.
Several methods have been used to remove synthetic dyes, such as biodegradation,
photocatalytic degradation, oxidation processes, membrane filtration, and adsorption. 5,6
Among these, adsorption using activated carbon (AC) is the most popular because of its
effectiveness and low cost. 7 To enhance the adsorption capacity of AC for removing toxic
contaminants, various physical, chemical, and biological treatment techniques have been
applied, 8 in which the generation of surface functional groups always plays a significant role
in the adsorption process. It has been discovered that the chemical modification of AC,
including by acid or alkali treatment, can greatly improve the efficiency of dye and heavy
metals removal compared to other treatment methods. 9,10 However, chemical treatment
using highly corrosive acid or alkaline hydroxides produces a large amount of secondary
wastewater in the post-treatment process. In response to such a problem, a simple thermal
air oxidation method has been developed, in which oxygenated functional groups can be
generated on the surface of AC, which is beneficial for the adsorption performance. 11,12 In
particular, it provides an environmentally friendly and low-cost means of enhancing the
adsorption capacity of AC toward contaminant removal.
Nevertheless, the utilization of AC has a limitation in its repeated use for dye
adsorption. 13 In addition, the toxicity of the adsorbed dyes on the AC is a problem as these
cannot decompose and reduce the toxicity. 14 The regeneration processes applied for the
spent AC from MB adsorption includes electrochemical, 15 microwave, 16 high-temperature
treatment under a special atmosphere (N2, 17 steam, 18 or CO2 (ref. 19 )), and thermal
oxidation treatment. 19 Among these, the most feasible and widely accepted for industrial
applications is thermal air oxidation regeneration. However, there is no report on the
behavior of this method during the regeneration process of spent AC and no details
confirming that all the MB molecules are decomposed on AC. Therefore, it is necessary to
:
investigate the thermal oxidation on spent AC by using a thermogravimetric analyzer coupled
with a differential thermal analyzer (TGA-DTA), and to clarify the degradation of MB before
and a"er the regeneration process utilizing a carbon, hydrogen, nitrogen, sulfur (CHNS)
analyzer. In addition, reusability of the thermal air-regenerated adsorbent was observed for
at least three cycles, whereas it was previously reported that only one cycle was possible. 19
In this study, a thermal air oxidation process was applied to regenerate the spent AC to
recover the adsorption performance and a pre-treatment process could increase the
adsorption capacity. Meanwhile, this process was expected to be able to produce more
oxygenated functional groups on the regenerated AC to enhance the adsorption capacity, as
reported in the literature. 11 The physicochemical properties of the modified adsorbent were
characterized by Boehm titration and nitrogen adsorption measurements. Furthermore, the
key parameters, such as kinetics, equilibrium, and thermodynamics, for MB adsorption were
investigated.

Experimental

Modification of the AC surface by thermal air oxidation treatment

This study used commercial AC, such as Haycarb AKO 8 × 30 (Haycarb PLC, Sri Lanka),
produced from steam-activated carbon coconut shell (SACCS), 20 as a starting material for the
adsorbent. First, the SACCS was ground and sieved to obtain a particle size of 0.1–0.3 mm,
a"er which the surface of the SACCS was modified by thermal air oxidation treatment. Briefly,
1 g of SACCS was spread out on a porcelain ceramic crucible with a dimension of 120 × 60
mm, and then put in to a muffle furnace (Nabertherm B180, Nabertherm GmbH, Germany),
and heated at a heating rate of 5 °C min−1 to the target final temperature (i.e., 300 °C, 325 °C,
350 °C, 375 °C, and 400 °C, respectively), and then maintained at this final temperature for 1 h.
A"erward, the sample was cooled to room temperature, and the obtained air-oxidized SACCS
was denoted as SACCS-X, where X represents the target temperature.

Characterizations

The yield from the thermal air oxidation treatment of the SACCS was calculated as the
:
percentage of the mass of air-oxidized SACCS (SACCS-X) relative to the raw material. The
amount of weakly acid functional groups in the SACCS-X was characterized by Boehm's
titration method with different degrees of 0.01 N bases (NaOH, NaHCO3, and Na2CO3). 21
Furthermore, the BET surface area and pore-size distribution of the adsorbent were
determined through nitrogen adsorption measurements at 77 K using a surface area and
pore-size analyzer (NOVA 4200e, Quantachrome Instrument Corp., USA). Degassing treatment
was applied at 150 °C and 0.1 mbar before the nitrogen adsorption measurements. The
surface areas of the SACCS and SACCS-X were determined in the range of P/P0 = 0.01–0.30.

Additionally, elemental analysis of the adsorbent was performed with a CHNS analyzer (Vario
EL Cube, Elementar Analysensysteme GmbH, Germany). The thermal air oxidation behavior
on both the fresh and spent SACCS and SACCS-350 was characterized by TGA-DTA (STA200RV,
Hitachi Co., Ltd., Japan).

Adsorption experiment

The adsorption of methylene blue (MB) (Merck Millipore Corp., Singapore) on the SACCS and
SACCS-350 were observed by adding 15 mg of adsorbent and 25 mL of MB solution (100 mg
L−1) into a 50 mL centrifugation tube. The adsorption experiments were conducted on a multi-
plate vortex (Joan Lab Equipment Co., Ltd., China) in an incubator chamber for 3 h with a
shaker speed and temperature of 500 rpm and 30 °C, respectively. A"er reaching equilibrium,
the suspension was centrifuged, and the solution was analyzed by UV-Vis spectrophotometer
(Genesys 10S UV-Vis, Thermo Scientific Corp., USA) at a wavelength of 664 nm. The
adsorption of each batch was performed at least 2 times, and the following equation was
used to calculate the amount of adsorbed MB at equilibrium (qe, mg g−1):

(1)

where Ce is the equilibrium concentration of the adsorbate (mg L−1), C0 is the initial
concentration of the adsorbate (mg L−1), V is the volume of solution (L), and m is the mass of
the adsorbent (g).
:
Adsorption kinetics studies

The study of the kinetics model for MB adsorption on the SACCS or SACCS-350 was conducted
in a similar way as the adsorption experiments in the time range of 10–360 min. The
adsorption kinetics was determined with pseudo-first-order (PFO), pseudo-second-order
(PSO), and intraparticle diffusion Weber–Morris (ID-WM) models. Table S1 (eqn (S1)–(S3)) †
shows the list of the mathematical equations for the kinetics models.

Adsorption isotherm studies

The study of the isotherm model for MB adsorption on the SACCS or SACCS-350 was
conducted as described in the Adsorption experiment section, with different initial
concentrations of MB (25, 50, 100, 200, 300, and 400 mg L−1) for 3 h. The Langmuir, Freundlich,
Redlich–Peterson, and Temkin models were used for the adsorption isotherm studies. Table
S1 (eqn (S4)–(S7)) † presents the list of the mathematical equations in the isotherm models.

Determination of the thermodynamic parameters

The thermodynamic parameters for MB adsorption on the SACCS or SACCS-350 were

determined as described in the Adsorption experiment section, with different temperatures
of 30 °C, 40 °C, 50 °C, and 55 °C for 3 h.

Normalized adsorption efficiency (NAE)

The normalized ratio of the adsorption efficiency was used to describe the visibility of the
yield of SACCS-X toward the adsorption performance of SACCS-X compared to the starting
material (SACCS). Therefore, the following was used to calculate the NAE:

(2)

where qe,SACCS and qe,SACCS-X are the amount of adsorbed MB on SACCS and SACCS-X at
equilibrium (mg g−1), respectively, and yieldSACCS-X is the yield of SACCS-X a"er thermal air
oxidation treatment (%).
:
Results and discussion

Characterizations of SACCS and SACCS-X

In this study, steam-activated carbon coconut shell (SACCS) was used rather than chemical
activated carbon in order to avoid the chemical impurities from the activation process, which
would otherwise contribute to the MB adsorption. In the preliminary step, the temperature
effect on the thermal air oxidation of SACCS toward the adsorption performance was studied
in the temperature range of 300–400 °C for 1 h. Fig. S1 † shows that the adsorbent yield
decreased when the treatment temperature increased. TGA-DTA analysis was used to monitor
the thermal air oxidation process of SACCS (Fig. S2a † ), showing that the mass profile (TG)
started to decrease from 150 °C and simultaneously the DTA profile exothermically increased,
which corresponded to the oxidation process of SACCS (Fig. S2a † ). Among those obtained at
different temperatures through air oxidation, SACCS-400 showed the highest mass loss of 89.6
wt% due to the overoxidation (Fig. S1 † ). Therefore, it was not considered for the adsorption
of MB due to the ineffective use of this adsorbent with a mass loss of more than 50 wt%.
The adsorption efficiencies (qe) of SACCS and SACCS-X were evaluated for MB adsorption
at 30 °C for 3 h. As summarized in Fig. 1 , the adsorption efficiency of the SACCS-X increased
with the elevation of the thermal air oxidation temperature. Herein, the enhanced
performance resulted from the enriched total oxygenated functional groups (phenolic,
lactonic, and carboxylic groups) on the surface of SACCS-X, as confirmed by Boehm titration,
as shown in Table 1 and Fig. S3. † This was because they could improve the adsorption
performance of the adsorbent due to electrostatic and hydrogen bond interactions. 11,22,23
Meanwhile, the normalized adsorption efficiency (NAE) ( eqn (2) , Fig. 1 ) was further used to
evaluate the adsorption performance. SACCS-375 showed the highest adsorption capacity,
corresponding to the greatest amount of carboxylic groups with the lowest adsorption energy
value compared to phenolic and lactonic groups. 24 However, despite having the highest
adsorption capacity, the NAE of SACCS-375 was lower than that of SACCS-350, which should
be related to other surface properties. Table 1 shows that the surface area of SACCS increased
from 467.9 to 688.0 m2 g−1 with the rise in air oxidation temperature until 350 °C, but then
:
decreased to 621.8 m2 g−1 at 375 °C. A similar variation pattern also occurred on the
cumulative pore volumes of SACCS-X. As a result, since SACCS-350 had a higher surface area
and a larger cumulative pore volume, more MB molecules should be adsorbed on its
surface, 22,25,26 corresponding to the pore-size distribution, as shown in Fig. S4. † The
increase in oxidation temperature resulted in the expansion of the SACCS half-pore width
from 9–11 Å to 12–21 Å. However, treatment at over 350 °C led to the collapse of the
adsorbent pore properties due to overoxidation, as indicated by the decreasing surface area
and cumulative pore size of SACCS-375. These results also proved that SACCS-350 should
have a higher normalized adsorption efficiency than SACCS-375. Therefore, SACCS-350 was
selected as the optimum SACCS-X absorbent for the further equilibrium, kinetic,
thermodynamic, and reusability studies on MB adsorption.

Fig. 1 Adsorption efficiency (qe) and normalized adsorption efficiency from SACCS and
SACCS-X. Adsorbent = 15 mg, MB = 100 mg L−1, 500 rpm, 30 °C, 180 min.

Table 1 Surface characteristics of SACCS and SACCS-X

:
 b
Adsorbent SBET a (m2 g−1) Cumulative pore volume (cm g ) Total oxygenated functional groups
3 −1

SACCS 467.91 0.274 557.77

SACCS-300 571.57 0.341 668.66
SACCS-325 615.39 0.359 896.41
SACCS-350 688.02 0.408 1185.26
SACCS-375 621.77 0.370 1257.49

a SBET was calculated using the N2 isotherm data from the multiple point method within a P/P low

Adsorption kinetics

The adsorption kinetics was used to describe adsorbate removal on the adsorbent, which is
beneficial to understand the mechanism of the adsorption process. This study applied the
PFO and PSO models to fit the kinetics data to explain the adsorption process, as presented in
Table S1, eqn (S1)–(S3). † Fig. 2a and Table 2 show the adsorption kinetics of MB on SACCS
and SACCS-350, and it could be observed that the adsorption profiles of MB on both SACCS
and SACCS-350 increased with the extension of time. A rapid adsorption of MB occurred in the
initial step, corresponding to a higher occupancy of adsorption sites for MB molecules.
Subsequently, the rate of adsorption decreased at 60 min since a large amount of MB
molecules were already attached to the adsorption sites, and then it reached equilibrium at
180 min. As indicated in Table 2 , the PSO model fitted the adsorption kinetics for both SACCS
and SACCS-350, with the values of R2 of 0.9794 and 0.9925, respectively. The calculated qe
values based on the PSO model were also close to the values for the experimental qe. These
results indicated that MB adsorption on SACCS and SACCS-350 should occur in the
chemisorption process, in which those oxygenated functional groups on SACCS and SACCS-
350 should effectively contribute to anchoring the MB molecules.
:
 Fig. 2 Adsorption kinetics of MB on SACCS and SACCS-350. (a) PFO and PSO kinetics models
and (b) intraparticle diffusion of the Weber–Morris kinetics model. Adsorbent = 15 mg, MB =
100 mg L−1, 500 rpm, 30 °C.

Table 2 Kinetic parameters of the PFO, PSO, and intraparticle diffusion of the Weber–Morris
(ID-WM) models for the adsorption of MB on SACCS and SACCS-350
Models Parameters SACCS SACCS-350
Pseudo-first-order model (PFO) qe,exp (mg g−1) 29.05 110.08

qe (mg g−1) 13.51 34.31

k1 (min−1) 0.0017 0.0070

R2 0.8876 0.9797

Pseudo-second-order model (PSO) qe (mg g−1) 36.65 129.75

:
 k2 (mg g−1 min−1) 0.0007 0.0002

R2 0.9794 0.9925
Intraparticle diffusion of Weber–Morris model (ID-WM) kip1 (mg g−1 min1/2) 4.8908 13.6524

kip2 (mg g−1 min1/2) 2.0189 6.1017

kip3 (mg g−1 min1/2) 0.7414 1.5490

The adsorption kinetics processes of SACCS and SACCS-350 occurred following the PSO
model. However, the PSO model was unsuitable for the mechanistic analysis. 27 As a result,
the Weber–Morrison (ID-WM) model, which describes intraparticle diffusion, was applied to
determine the reaction control step and the adsorption mechanism. Fig. 2b shows the
multilinearity plots of qt versus t1/2, corresponding to the adsorptions of MB on both SACCS

and SACCS-350 occurring in multiple steps of intraparticle diffusion. 28 Herein, the

parameters of the intraparticle diffusion model for SACCS-350 were 13.6524, 6.1017, and
1.5490 mg g−1 min1/2, respectively, which were higher than those of SACCS of 4.8908, 2.0189,
and 0.7414 mg g−1 min1/2. These results indicated that the modification of SACCS with thermal
air oxidation at 350 °C could increase the rate of MB adsorption, thereby improving the
adsorption capacity. This indicated that the adsorption kinetics of MB on either SACCS or
SACCS-350 should be controlled by chemisorption and intraparticle diffusion.

Adsorption isotherms

Isothermal adsorption is always used to determine the equilibrium relationship between the
concentration of an adsorbate and the amount of accumulated adsorbate on the adsorbent.
Here, the adsorption isotherm results were fitted with four models, namely the Langmuir,
Freundlich, Redlich–Peterson, and Temkin models. The results of the fitting calculations are
summarized in Fig. 3 and Table 3 , while the linear and non-linear equations are listed in
Table S1 (eqn (S4)–(S7)). † The order of the adsorption isotherm fitting models for SACCS was
observed to be Redlich–Peterson > Langmuir > Freundlich > Temkin, while for SACCS-350, it
was Freundlich > Langmuir > Temkin > Redlich–Peterson.
:
 Fig. 3 Adsorption isotherms of SACCS and SACCS-350. Adsorbent = 15 mg, 500 rpm, 30 °C,
180 min.

Table 3 Adsorption isotherm parameters of MB on SACCS and SACCS-350

Models Parameters SACCS SACCS-350
Langmuir qmax (mg g−1) 79.31 160.36

KL (L mg−1) 0.0121 0.0875

R2 0.9283 0.9860

Freundlich KF (mg g−1(L mg−1)1/n) 6.49 55.34

1/n 0.3912 0.1856

R2 0.9264 0.9908

Redlich–Peterson (R–P) KRP (L g−1) 271.63 1218.91

αRP (L mg−1) 49.70 20.21

β 0.5721 0.8324

R2 0.9842 0.8645
Temkin KT (L g−1) 0.1867 36.6021

Hads (J mol−1) 169.07 155.35

R2 0.8726 0.9705
:
 The Langmuir isotherm assumes monolayer adsorption onto the homogeneous surface,
where the adsorption sites have the same energy with no adsorbate interaction in the surface
plane. 29,30 However, the Freundlich isotherm is applied to describe the adsorption of
heterogeneous surfaces with various affinities of adsorption sites. 31 The results showed that
SACCS-350 could be well-fitted to the Freundlich isotherm model with an R2 value of 0.9908,
as shown in Table 3 . This indicated the occurrence of different chemical interactions of the
adsorbate with the heterogeneous oxygenated functional groups on the surface of SACCS-
350. Meanwhile, the adsorption equilibrium of MB on SACCS was between the Langmuir and
Freundlich models with R2 values of 0.9283 and 0.9264, respectively, as demonstrated in
Table 3 . Therefore, the Redlich–Peterson model could be applied to describe MB adsorption
on the SACCS adsorbent, which includes features of the Langmuir and Freundlich isotherms,
and represents the adsorption equilibrium over a broad range of concentrations of the
adsorbate. 32 The exponent, β, lay between 0 and 1. The Redlich–Peterson and Langmuir
equations were the same when β = 1, but, at β = 0, it followed Henry's law. Since the Redlich–
Peterson equation was related to SACCS with a β value of 0.5721 and R2 of 0.9842, it indicated

that the behavior of the equilibrium isotherm of MB was between the Langmuir and
Freundlich isotherms.
In addition, the Temkin isotherm model was also used to describe the MB adsorption
equilibrium on SACCS and SACCS-350, in which the heat of adsorption of all the molecules on
the layer was assumed to decrease linearly with the coverage due to the adsorbent–
adsorbate interactions with a uniform distribution of binding energies. 33 Herein, the positive
values of Hads for SACCS and SACCS-350 indicated that adsorption occurred in the

endothermic process ( Table 3 ). However, this model had a lower agreement for SACCS (R2 =
0.8726) and SACCS-350 (R2 = 0.9705) than the Redlich–Peterson model (R2 = 0.9842) for SACCS

and Freundlich model (R2 = 0.9908) for SACCS-350, indicating that the variation in adsorption
heat did not conform to the assumed linear decrease.

Adsorption thermodynamics

The thermodynamic behavior was studied to determine the change of enthalpy (ΔH), entropy
:
(ΔS), and Gibbs energy (ΔG) by considering the unit transfer of the adsorbate from the
solution onto the liquid–solid interface. The thermodynamic parameters were calculated
using the following equations: 34,35

(3)

ΔG = −RT ln Kd (4)

ΔG = ΔH − TΔS (5)

(6)

where R (8.314 J K−1 mol−1) is the universal gas constant, T (K) is the absolute solution
temperature, and Kd is the distribution coefficient.

Table 4 shows the thermodynamic parameters for the adsorption of MB on SACCS and
SACCS-350 in the temperature range of 30–55 °C (303–328 K), which were determined based
on the results shown in Fig. S5. † It was observed that the increase in adsorption temperature
resulted in a decrease in the ΔG value for both SACCS and SACCS-350. The positive value of ΔG
for SACCS (ΔG = 2.62 to 0.48) implied that it was more difficult for MB adsorption to occur on
the SACCS's surface. In contrast, the value of ΔG for SACCS-350 was negative, ranging from
−2.99 to −6.01. This indicated that the adsorption process could occur naturally due to the
thermal oxidation modification of the adsorbent surface. 36 The positive ΔS values of both
SACCS and SACCS-350 at 80.35 and 120.15 J K−1 mol−1, respectively, were attributed to the

high disorder and randomness on the liquid–solid interface during MB adsorption.

Meanwhile, the ΔH values of both SACCS and SACCS-350 were also positive, indicating that
MB adsorption was an endothermic process. 37

Table 4 Thermodynamics parameters of SACCS and SACCS-350 on MB adsorption. Adsorbent

:
= 15 mg, MB = 100 mg L−1, 500 rpm, 180 min

ΔG (kJ mol−1)
Adsorbent ΔH (kJ mol−1) ΔS (J K−1 mol−1) 303 K 313 K 323 K 328 K R2
SACCS 26.99 80.35 2.62 1.80 1.23 0.48 0.9778
SACCS-350 33.39 120.15 −2.99 −4.24 −5.40 −6.01 0.9996

Comparison of regeneration methods for spent AC

Table 5 compares the regeneration methods for spent activated carbon (AC) in MB
adsorption. Among them, the microwave technique has the shortest time (0.05 h) to
regenerate spent AC. The high equipment price for large-scale processes limits its
implementation on the industrial scale. Furthermore, the adsorption capacity of regenerated
AC decreases with the increasing adsorption–regeneration cycles due to the deterioration of
the pore structure of AC. 16

Table 5 Comparison of regeneration method for spent activated carbon in MB adsorption

No. Activated carbon sources SSA (m2 g−1)

1 Bituminous coal (commercial) 862
PS: 0.55–0.75 mm
AM: n.a.
2 Coconuts shells 1726
PS: 0.25–0.425 mm
AM: NaOH
3 Spent coal (commercial) 1233
PS: 2–3 mm
AM: n.a.
4 Charcoal (commercial) 651 (SAC) f ; 768 (Pyro-RAC) g; 730 (CO2-RAC) h
PS: 0.3 mm
AM: n.a.
5 Coconut shells (commercial) 1131
:
 PS: 1 mm
AM: steam
6 Coconut shells (commercial) 467.91 (SACCS); 688.02 (SACCS-350)
PS: 0.1–0.3 mm
AM: steam and thermal

a Maximum monolayer adsorption capacity. b Initial concentration of MB for qmax determination.

AM: activation method; n.a.: not available.

The high regeneration temperature of spent AC in N2, CO2, and steam conditions has also

been studied, as presented in Table 5 . However, there is no reusability data for more than 1
cycle. Also, the need for high-temperature treatment conditions impacts the high cost of the
process. 17–19
Unlike the above-mentioned regeneration processes for spent AC, the thermal air
oxidation method does not require special equipment, and the treatment temperature is
relatively low. Therefore, it represents a reasonable regeneration process for spent AC
compared to the other approaches.

Evaluation of the adsorbent reusability and feasibility of thermal air oxidation

treatment

Although the thermal oxidation method has been reported for the regeneration of spent
AC, 19 there is no information regarding the reusability of thermal air-regenerated adsorbents
in more than one cycle. In this study, the reusability of the adsorbent was evaluated on
SACCS-350 since it exhibited the highest adsorption capacity among all the SACCS and the
SACCS-X. Also, studies were conducted without and with the regeneration process. Fig. S6 †
shows the reusability of SACCS-350 without regeneration, in which the adsorption capacity
decreased from 110.08 to 44.0.6 mg g−1. This indicated that most of the adsorption sites of the

fresh SACCS-350 were occupied with MB molecules in the first adsorption process. When the
SACCS-350 was reused without regeneration, the adsorption capacity decreased due to the
reduction in available adsorption sites. The thermal oxidation treatment of the spent SACCS-
350 was performed to solve this problem. Fig. 4 shows that with the regeneration, the
:
adsorption ability of the SACCS-350 could be completely recovered even a"er being reused
and regenerated 3 times. To clarify the regeneration process, TGA-DTA analysis was used to
observe the decomposition of the adsorbed MB. Fig. S2b † shows the differences between the
TGA profiles of fresh and the regenerated SACCS-350. This indicated that the adsorbed MB
could be completely decomposed with thermal oxidation treatment. Meanwhile, the
decrease in the DTA profile from 200 °C to 500 °C should be related to the endothermic
reaction due to the decomposition of MB. Despite the MB molecules decomposing, the
nitrogen element remained and accumulated on the SACCS-350 even a"er four times of the
regeneration process ( Fig. 5 ). It should be noted that the accumulated nitrogen on the
adsorbent a"er the regeneration process did not affect the adsorption capacity of the
regenerated SACCS-350, hence, it could be reused for 3 times or more.

Fig. 4 Reusability of SACCS-350 with a regeneration process at 350 °C. Adsorbent = 15 mg,
MB = 100 mg L−1, 500 rpm, 30 °C, 180 min.
:
 Fig. 5 CHNS analysis of SACCS-350 before and a"er 4 times regeneration with the thermal
air oxidation method at 350 °C for 1 h. The inset shows the enlargement of H, N, and S mass
fraction elements.

Conclusions

Thermal air oxidation treatment was successfully used to increase the adsorption capacity of
AC in MB removal and to regenerate spent AC. It was observed that SACCS-350 had the
highest normalized adsorption efficiency (NAE). This was increased by 2.8 times compared to
that of the non-treated SACCS. The adsorption kinetic process for SACCS and SACCS-350
followed the PSO model. Meanwhile, the equilibrium data of SACCS and SACCS-350 on MB
adsorption showed good agreement with the Redlich–Peterson and Freundlich isotherm
models, respectively. The thermodynamic study indicated that the thermal air oxidation
treatment could make the endothermic adsorption processes more spontaneous.
Furthermore, the thermal air oxidation treatment could recover the adsorption ability of the
spent SACCS-350 at least 3 times. Therefore, this study can be expected to provide guidance
to improve the adsorption effectiveness of activated carbon and to regenerate the spent AC
so that it can be reused, resulting in a more environmentally friendly process and lower price.

Author contributions
:
Irwan Kurnia: conceptualization, methodology, investigation, writing – original dra", project
administration, supervision, resources. Surachai Karnjanakom: validation, writing – review &
editing. Irkham Irkham: investigation, writing – review & editing. Haryono Haryono:
investigation, writing – review & editing. Yohanes Andre Situmorang: investigation, writing –
review & editing. Antonius Indarto: investigation, writing – review & editing. Atiek Rostika
Noviyanti: investigation, writing – review & editing, resources. Yeni Wahyuni Hartati:
investigation, writing – review & editing, resources. Guoqing Guan: validation, writing –
review & editing.

Conflicts of interest

There are no conflicts to declare.

Acknowledgements

This work is supported by Research Funding of Associate Professor Acceleration (Hibah RPLK)
and Academic Leadership Grant No. 2203/UN6.3.1/PT.00/2022 from Universitas Padjadjaran.

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:
 Footnote

† Electronic supplementary information (ESI) available. See DOI:

https://doi.org/10.1039/d2ra06481b

This journal is © The Royal Society of Chemistry 2023

:
 Chemical Engineering Journal 219 (2013) 499–511

Contents lists available at SciVerse ScienceDirect

Chemical Engineering Journal

journal homepage: www.elsevier.com/locate/cej

Review

An overview of the modification methods of activated carbon for its

water treatment applications
Amit Bhatnagar a,⇑, William Hogland a, Marcia Marques a,b, Mika Sillanpää c
a
School of Natural Sciences, Linnaeus University, SE-391 82 Kalmar, Sweden
b
Department of Sanitary and Environmental Engineering, Rio de Janeiro State University, UERJ, Rio de Janeiro, Brazil
c
Faculty of Technology, Lappeenranta University of Technology, Patteristonkatu 1, FI-50100 Mikkeli, Finland

h i g h l i g h t s

" Various modification methods of activated carbon (AC) have been reviewed.
" Application of modified carbons in water treatment has been reviewed.
" Comparison of virgin and modified AC in pollutants removal has been discussed.
" Future perspectives in this direction have been proposed.

a r t i c l e i n f o a b s t r a c t

Article history: Activated carbon has been recognized as one of the oldest and widely used adsorbent for the water and
Received 27 August 2012 wastewater treatment for removing organic and inorganic pollutants. The application of activated carbon
Received in revised form 9 December 2012 in adsorption process is mainly depends on the surface chemistry and pore structure of porous carbons.
Accepted 10 December 2012
The method of activation and the nature of precursor used greatly influences surface functional groups
Available online 20 December 2012
and pore structure of the activated carbon. Therefore, the main focus of researchers is to develop or mod-
ifies the activation/treatment techniques in an optimal manner using appropriate precursors for specific
Keywords:
pollutants. In recent years, emphasis is given to prepare the surface modified carbons using different pro-
Aquatic pollutants
Water treatment
cedures to enhance the potential of activated carbon for specific contaminants. Various methods such as,
Activated carbon acid treatment, base treatment, impregnation treatment, ozone treatment, surfactant treatment, plasma
Surface modification treatment and microwave treatment have been studied to develop surface modified activated carbons. In
Modification methods this paper, these modification methods have been reviewed and the potential of surface modified acti-
vated carbons towards water treatment has been discussed. This review article is aimed at providing pre-
cise information on efforts made by various researchers in the field of surface modification of activated
carbon for water pollution control.
Ó 2012 Elsevier B.V. All rights reserved.

Contents

1. Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 500
2. Methods for the modification of activated carbon . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 500
2.1. Acidic treatment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 500
2.2. Base treatment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 502
2.3. Impregnation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 504
2.4. Microwave treatment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 505
2.5. Ozone treatment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 505
2.6. Plasma treatment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 506
2.7. Biological modification . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 507
2.8. Miscellaneous modification methods of AC. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 507

⇑ Corresponding author. Tel.: +46 (0) 480 44 7352.

E-mail addresses: amit.bhatnagar@lnu.se, dr.amit10@gmail.com (A. Bhatnagar).

1385-8947/$ - see front matter Ó 2012 Elsevier B.V. All rights reserved.
http://dx.doi.org/10.1016/j.cej.2012.12.038
500 A. Bhatnagar et al. / Chemical Engineering Journal 219 (2013) 499–511

3. Cost estimation. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 508

4. Concluding remarks . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 509
Acknowledgments . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 509
References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 509

1. Introduction ment methods that enable AC surface for higher uptake of specific
pollutants. In general, the surface modification of AC is carried out
Activated carbon (AC) has been recognized as one of the most after the activation step. The modification can be classified into
popular and widely used adsorbent in water and wastewater treat- three categories: (i) chemical modification, (ii) physical modifica-
ment throughout the world [1–5]. Charcoal, the forerunner of mod- tion and, (iii) biological modification. Furthermore, oxidative [18]
ern activated carbon, is the oldest adsorbent known in water and non-oxidative [19] methods of surface treatments of AC have
purification. The specific adsorptive properties of charcoal were been reported in the literature. In previous studies, the presence
first discovered by Scheele in 1773 for the treatment of gases fol- of complexing groups on AC surface such as, carboxylic, lactonic,
lowed by decolorizing of solutions in 1786 and he provided the and phenolic groups have been widely examined [20,21]. Before
first systematic account of the adsorptive power of charcoal in performing any modification to AC, it is important to understand
the liquid phase [6]. In the following years, Lowitz established various factors which influence the adsorption potential of AC. It
the use of charcoal for the removal of bad taste and odors from will help to tailor their specific physical and chemical characteris-
water during 1789 to 1790. The credit to develop commercial acti- tics to enhance its affinities towards different pollutants present in
vated carbon goes to a Swedish chemist von Ostreijko who ob- aqueous solution. In this review, different modification methods,
tained two patents, in 1900 and 1901, on covering the basic which have been applied to AC, have been reviewed and their main
concepts of chemical and thermal (or physical) activation of car- outcomes have been discussed. For information pertaining to de-
bon, with metal chlorides and with carbon dioxide and steam, tailed experimental methodology and conditions, readers are re-
respectively [6]. The process of chemical activation of sawdust ferred to the full articles listed in the references.
with zinc chloride was carried out for the first time in an Austrian In this review, the progress made in the last few years concern-
plant at Aussing in 1914 on an industrial scale, and also in the dye ing the synthesis of surface modified ACs is summarized and de-
plant of Bayer in 1915 [7]. In this type of activation, pyrolytical tails of different methods have been discussed. Also, the
heating of the carbonaceous material was performed in the pres- application of surface modified ACs in water treatment has also
ence of dehydrating chemicals such as, zinc chloride or phosphoric been presented by comparing the properties between crude and
acid [8]. modified active carbons. Additionally, some information has been
Activated carbon is a common term used to describe carbon- presented for the treatments which only look for the improvement
based materials which contains well developed internal pore struc- of the mechanical properties of the activated carbons and so, in-
ture. AC is produced from a variety of carbonaceous rich materials crease the life of the active carbon. Finally, the future perspectives
such as wood, coal, lignite and coconut shell [9]. The high surface of the research in this field are highlighted.
area, large porosity, well developed internal pore structure consist-
ing of micro-, meso- and macropores as well as a wide spectrum of 2. Methods for the modification of activated carbon
functional groups present on the surface of AC make it a versatile
material which has numerous applications in many areas, but The modification of the surface chemistry of carbons is consid-
mainly in the environmental field. ered as a promising and attractive way toward new applications of
Oxygen, hydrogen, sulfur and nitrogen are generally present in carbon in many fields. Modification of AC involves oxidation and
AC in the form of functional groups and/or atoms chemically further grafting onto the AC surface by chemical, electrochemical,
bonded to the structure. In the carbon structure, the main func- plasma and/or microwave method to introduce functional groups
tional groups which, in general, are considered to be responsible (e.g., carboxylic acid, amine, etc.) and molecules such as cyclodex-
for uptake of pollutants include carboxyl, carbonyl, phenols, lac- trin. Oxidation can be attained by chemical modification [22,23],
tones, quinones besides others. The unique adsorption properties air oxidation [23,24], electrochemical oxidation [25], and plasma
of AC can be significantly influenced by these functional groups. [26] or ozone treatment [27]. A detailed description of some
The functional groups on carbon surface mainly derived from acti- important methods used for the surface modification of AC during
vation process, precursor(s), thermal treatment and post chemical last decades has been presented in subsequent sections.
treatment. The nature and concentration of surface functional
groups may be modified by suitable thermal or chemical treat- 2.1. Acidic treatment
ments to improve the performance of AC for specific contaminants
removal. Acid treatment of carbon is generally employed to oxidize the
The efficiency of ACs as adsorbents for diverse types of pollu- porous carbon surface as it increases the acidic property, removes
tants is well reported [9,10]. It is well known that activated carbon the mineral elements and improves the hydrophilic nature of sur-
has been found much efficient for removing organic compounds face [28]. The nitric acid and sulfuric acid are the most widely stud-
than metals and other inorganic pollutants. Efforts are ongoing to ied ones besides some other acids used for the purpose. Acidic
substantially improve the potential of carbon surface by using dif- functional groups (i.e., oxygen functional groups containing proton
ferent chemicals or suitable treatment methods [11] which will en- donors) on carbon surfaces have been examined for the removal of
able AC to enhance its potential for the removal of specific heavy metals from water and found to be highly favorable because
contaminants from aqueous phase. The physical and chemical metal ions have a tendency to form metal complexes with the neg-
structure of carbon could be changed by various methods, i.e. acti- atively charged acid groups. Aggarwal et al. [29] prepared two
vation conditions (different agents, temperature and time of the samples of modified granular activated carbons and two samples
process), precursor, additives, etc. [12]. Different methods have of fibrous activated carbons by oxidizing them with nitric acid,
been reported in the literature [13–17] to modify AC surface which ammonium persulphate and hydrogen peroxide in the solution
throw some light on the chemistry and mechanism behind treat- phase and with gaseous oxygen at 350 °C to enhance the amount
 A. Bhatnagar et al. / Chemical Engineering Journal 219 (2013) 499–511 501

of carbon–oxygen surface chemical structures. The activated car- apparent from the enthalpies of adsorption for various carbons,
bons were also degassed at different temperatures between which varied markedly with adsorbate concentration and surface
400 °C and 950 °C to gradually eliminate these surface chemical functional group concentration resulting from the heat treatment.
structures. It was observed that the adsorption of Cr(III) increased The effect of varying physical and chemical properties of acti-
on oxidation and decreased on degassing. On the other hand, the vated carbons on adsorption of elemental mercury (Hg0) has been
adsorption of Cr(VI) decreased on oxidation and increased on studied by Li et al. [32]. Heat treatment (1200 K) in nitrogen (N2),
degassing. air oxidation (693 K), and nitric acid (6 N) treatment of two bitumi-
A saturated solution of (NH4)2S2O8 in 1 M H2SO4 was used to nous coal-based activated carbons (BPL and WPL, both from Calgon
oxidize two commercial carbon materials [30]. It was noticed by Carbon) were conducted to vary their surface oxygen functional
the researchers that oxidation of as-received carbons greatly in- groups. Adsorption experiments were conducted using a fixed-
creased Zn(II) uptake under static conditions although decrease bed reactor at a temperature of 398 K under N2 atmosphere. Tem-
in surface area and carbon porosity after oxidation were observed. perature-programmed desorption (TPD) and base-acid titration
This was attributed to the fact that the oxidized carbon surface was experiments were conducted to determine the chemical character-
more negatively charged than that of the as-received carbons due istics of the carbon samples. Characterization of the physical and
to the dissociation of surface acid groups, which increased the elec- chemical properties of activated carbons in relation to their Hg0
trostatic adsorbate–adsorbent interactions. In this case, there was adsorption capacity provided important mechanistic insight on
a displacement of two H+ ions from the oxidized carbon surface Hg0 adsorption. Results suggested that oxygen surface complexes,
per Zn(II) ion adsorbed. It was also pointed out by the researchers possibly lactone and carbonyl groups, were the main active sites
that not all acid groups or even all carboxyl groups were involved for Hg0 uptake. The carbons which had a lower carbon monoxide
in the Zn(II) adsorption. This was because metal ions were ad- (CO)/CO2 ratio and a low phenol group concentration tend to have
sorbed on the external carbon surface and either electrostatically a higher Hg0 adsorption capacity, suggesting that phenol groups
repelled other Zn(II) ions in solution or block their accessibility might inhibit Hg0 adsorption. The high Hg0 adsorption capacity
to other surface acid groups located within internal pores of the of a carbon sample was also found to be associated with a low ratio
oxidized carbons. Adsorption of Zn(II) on non-oxidized, as-received of the phenol/carbonyl groups. A possible Hg0 adsorption mecha-
carbons with a high total surface basicity might result from Cp-cat- nism, which likely involved an electron transfer process during
ion interactions. Increased solution ionic strength reduced Zn(II) Hg0 adsorption was also discussed by the researchers which re-
uptake because of screening effect of the surface charge produced vealed that the carbon surfaces might act as an electrode for Hg0
by the added salt. In addition, an increase in the solution pH be- oxidation.
tween 3.0 and 6.0 resulted in an increase in Zn(II) uptake because Macias-Garcia et al. [33] used Merck carbon (1.5 mm) and trea-
of an increase in attractive adsorbate–adsorbent interactions. The ted it in three ways: heating from ambient temperature to 900 °C
oxidized activated carbon cloth was found to be much more effec- in SO2; treatment at ambient temperature in SO2; or successive
tive than the as-received activated carbon cloth for Zn(II) removal treatments in SO2 and H2S at ambient temperature. These samples
under dynamic conditions. The presence of tannic acid decreased were further tested as adsorbents for Cd2+ adsorption from aque-
the effectiveness of the oxidized activated carbon cloth bed to re- ous solution. It was found by the researchers that the amount of
move Zn(II). An increase in the initial concentration of tannic acid sulphur introduced was high when heated to 900 °C in SO2 and
had a greater effect on the removal of tannic acid than on Zn(II) by low when sulphurising also in SO2 but at an ambient temperature.
the column bed. This might be a result of the greater size of tannic A correlation was noted between the amount of sulphur intro-
acid molecules and their low affinity for oxidized carbon surfaces. duced and the loss of surface area and microporosity. The kinetics
Jia and Thomas [31] reported the incorporation of acidic oxygen of the adsorption process of Cd2+ was not markedly affected as a
functional groups into activated carbon by HNO3 oxidation and result of increased amount of sulphur in the samples. The adsorp-
studied the potential of modified carbon towards cadmium re- tion capacity was found to be dependent on the treatment applied
moval. It was observed that carboxylic acid groups were the major and also on the temperature and the solution pH. Heating to 900 °C
surface species incorporated and phenol and quinone groups were was found to be the most effective treatment, and the adsorption
introduced during the oxidation process. The formation of lactone in this case was 70.3%. The adsorption increased at 45 °C but de-
groups during heat treatment was also suggested by the workers. creased at pH 2.0 when compared to the adsorption at 25 °C and
The functional groups had a range of thermal stabilities with car- at pH 6.2, respectively. A low-cost activated carbon was prepared
boxylic acid groups being the least stable. Cadmium adsorption from apricot stone by chemical activation with sulphuric acid for
was not found to correlate with the changes in the porous struc- the adsorption of Pb(II) from aqueous solution [34]. The activated
tures of the activated carbons. However, cadmium adsorption carbon developed showed substantial capacity to adsorb Pb(II)
was dramatically enhanced after the carbon was oxidized and de- from dilute aqueous solutions. The percent removal was found to
creased significantly after heat treatment to progressively elimi- increase with solution pH from 1.5 to 6.0. The optimum pH was
nate the oxygen functional groups of various thermal stabilities, found to be 6.0 for maximum adsorption (88.3%) of Pb(II) from
thereby establishing a link between acidic oxygen functional the solution with initial lead concentration of 50 mg/L. The esti-
groups and cadmium adsorption, with the carboxylic acid groups mated adsorption capacity of apricot stone activated with sulphu-
contributing most to the cadmium adsorption. The ratios of H+/ ric acid for lead ions was 21.38 mg/g at an initial pH of 6.0.
Cd2+ for oxidized carbon and Na+/Cd2+ for Na+ ion exchange form A commercial activated carbon, Chemviron F 400, was also
of the carbon after adsorption were approximately 2, coinciding modified via oxidation with nitric acid to introduce a variety of
with the stoichiometry of cation exchange. Both reversible and acidic surface functional groups on carbon surface [35]. Results re-
irreversible adsorptions were reported for cadmium adsorption vealed that surface area and pore volume were reduced after oxi-
by oxidized carbon. The adsorption was found to be dominated dation treatment. However, the carbon surface acquired an acidic
by an irreversible process at relatively low equilibrium adsorbate character with carboxylic groups being the dominant surface func-
concentration, whereas reversible adsorption occurred at higher tional groups. The modified sample displayed cation-exchange
concentration when cadmium concentration was in excess of the properties over a wide range of pH values and exhibited polyfunc-
amounts of acidic groups. After heat treatment to eliminate car- tional nature. Both carbon samples were tested for the removal of
boxylic functional groups, the irreversible adsorption decreased copper(II), nickel(II), cobalt(II), zinc(II), and manganese(II). The
markedly. The differences in adsorption mechanism were also affinity series Mn2+ < Co2+ < Ni2+ < Cu2+ > Zn2+ was found to coin-
502 A. Bhatnagar et al. / Chemical Engineering Journal 219 (2013) 499–511

cide with the general stability sequence of metal complexes (the Ir- higher pH resulted in the negatively charged surface adsorbent
ving–Williams series). The higher preference displayed by carbons and caused higher adsorption.
towards copper(II) was a consequence of the fact that copper(II) of- The surface chemistry of a commercial activated carbon (Norit
ten forms distorted and more stable octahedral complexes. Liu ROX 0.8) with a slightly basic nature was modified by (i) oxidation
et al. [36] prepared modified coconut-based activated carbon by with HNO3 and (ii) thermal treatment under H2 flow at 700 °C in
treating it with HNO3 and NaOH and their potential was further as- order to obtain samples with acidic and basic properties, without
sessed for Cr(VI) removal. It was reported that nitric acid oxidation changing its textural parameters significantly [41]. The sorption
produced positive acid groups, and subsequently sodium hydrox- of a reactive dye, Rifafix Red 3BN (C.I. reactive red 241), on these
ide treatment replaced H+ of surface acid groups by Na+, and the three activated carbons was studied. The different uptake capaci-
acidity of activated carbon decreased. The coconut-based activated ties obtained on different samples were discussed in relation to
carbon was also pretreated with different concentrations of nitric the surface chemical properties of the adsorbents. It was found that
acid (from 0.5% to 67%) and was selected as palladium (Pd) catalyst the adsorption of the reactive (anionic) dye on the basic sample
support [37]. The Pd particle size and catalytic activity of Pd/C cat- (prepared by thermal treatment under H2 flow at 700 °C) was fa-
alysts were found to be highly dependent on the nitric acid concen- vored. This conclusion was explained on the basis of the dispersive
tration used in the pretreatment. Chen et al. [38] used citric acid to and electrostatic interactions involved.
modify commercially available activated carbon to enhance copper Besides dyes, the removal of phenols on modified ACs has also
adsorption from aqueous solutions. The carbon was modified with been examined. The adsorption of phenol onto activated carbon
1.0 M citric acid, followed by an optional step of reaction with Calgon F400 has been studied [42]. The carbon was modified by
1.0 M NaOH. It was found that the surface modification step re- acid treatment, using soxhlet extraction with 2 N HCl for 120 cy-
duced the specific surface area by 34% and point of zero charge cles. The treatment did not affect significantly the surface area of
(pHpzc) of the modified carbon by 0.5 units. Equilibrium results the activated carbon but affected the different functional groups,
showed that citric acid modification increased the adsorption and thus its adsorption properties. Tóth model reproduced satis-
capacity to 14.92 mg/g, which was 140% higher than the unmodi- factorily the experimental isotherm data and an adsorption enthal-
fied carbon. Higher initial solution pH resulted in higher copper py of 17.9 kJ/mol was found, which indicated that the process
adsorption. was exothermic. The pH also influenced the adsorption process
The adsorption of Hg(II) by a commercial activated carbon with and an empirical polynomial equation was able to reproduce max-
and without nitric acid treatment was studied in a batch system by imum capacity as a function of pH. The isotherms obtained at pH
Jazeyi and Kaghazchi [39]. Iodine adsorption test and nitrogen 3.0 and 7.0 were very similar and showed a higher adsorption
adsorption and desorption experiments were carried out to inves- capacity compared with that obtained at pH 13.0. The use of phos-
tigate the changes in porous characteristics during acid treatment. phate buffer solutions decreased the maximum phenol adsorption
FTIR analysis for both types of activated carbons was performed to capacity due to the competitive adsorption between the phenol
evaluate the effects of acid treatment on the surface functional and phosphates. Finally, it was demonstrated that the acid treat-
groups, which revealed oxidized surface in case of treated sample. ment introduced chloride ions into the carbon, giving it properties
Aqueous mercury equilibrium adsorption was enhanced by nitric of ion exchanger.
acid treatment by an average factor of two. This was interpreted As can be seen from the literature reviewed, acidic treatment of
by the formation of carboxylic acid surface groups, resulting in activated carbon was found to enhance the sorption of various pol-
an enhanced acid dissociation and chelating tendency with metal lutants onto modified forms of carbons due to the changes in the
ions. Percentage of adsorbed mercury at an equilibrium state in- surface chemistry. The acid treatments could remove the hydrox-
creased with increase in pH, and decreased with rise in initial mer- ide groups and produced a large number of oxygen-containing
cury concentration. The pH dependency was interpreted by the functional groups on the carbon surface. The number of acidic
changes of both mercury species and surface charge of activated functional groups is closely related to the capacity of activated car-
carbon at different pH. Furthermore, kinetic study using three- bon to adsorb metallic compounds [43].
stage model revealed the role of external and internal surfaces of
activated carbon in Hg(II) adsorption. Nitric acid treatment showed 2.2. Base treatment
enhanced adsorption of aqueous mercury on external and internal
surfaces of activated carbon by a factor of 2.87 and 2.09, Base (alkaline) treatment of AC produces positive surface charge
respectively. which in turn is helpful to adsorb negatively charged species in
Besides metal ions, the removal of dyes was also studied by higher amounts. The easiest way of producing porous carbons with
modified ACs. The effect of acidic treatments on activated carbons basic surface properties is to treat it at high temperature in inert,
for dye adsorption was investigated [40]. It was observed that hydrogen or ammonia atmosphere [44–49]. The treatment of AC
HNO3 treatment produced more active acidic surface groups such with NH3 at 400–900 °C leads to the formation of basic nitrogen
as carboxyl and lactone, resulting in the reduction in the adsorp- functionalities [50,51]. Amides, aromatic amines and protonated
tion of basic dyes. However, HCl treatment decreased active acidic amides are produced at 400–600 °C. Furthermore, pyridine-type
groups and thus enhanced the adsorption of larger dye molecules structures occur at higher temperatures, which enhance the basi-
on activated carbons. For methylene blue, the adsorption followed city of the carbon surface [52]. The nitrogen-containing groups
an order of AC > AC–HCl > AC–HNO3, while for crystal violet and generally provide the basic property, which could enhance the
rhodamine B, the adsorption order was: AC–HCl > AC > AC–HNO3. interaction between porous carbon and acid molecules, such as, di-
It was also found that solution pH showed a significant influence pole–dipole, H-bonding, covalent bonding, and so on. Furthermore,
on adsorption of methylene blue but little effect on rhodamine B. under alkaline (basic) solutions, it is expected that OH ions react
For methylene blue, higher pH resulted in higher adsorption. The with the surface functional groups of AC. The alkaline treatment of
adsorption increased from 4  104 to 14  104 mol/g at pH 3.0 AC is beneficial in enhancing the adsorption of especially organic
to 10.0. For rhodamine B, the adsorption at pH 3.0 and 5.0 was species (like phenol) from water.
much close and showed a slightly higher at pH 10.0. This different Modification of an activated carbon was performed using partial
adsorption behavior was because of the different functional groups oxygen gasification, nitric acid treatment, urea impregnation fol-
of dyes. Basically, methylene blue and other cationic dyes produce lowed by pyrolysis and pyrolysis in a urea saturated stream [53].
an intense molecular cation (C+) and reduced ions (CH+). Thus, It was reported that treatment of activated carbon by urea sup-
 A. Bhatnagar et al. / Chemical Engineering Journal 219 (2013) 499–511 503

ported the formation of basic groups and carbonyls. The presence er uptake of ammonia treated AC was attributed to the enlarged
of surface functional groups affected the adsorption capacity of carbon pores and basic surface formed during ammonia treatment.
the produced samples for the removal of phenols. Urea treated Recently, ammonia-modified activated carbon was prepared for
samples with a basic character and high nitrogen content pre- the adsorption of 2,4-dichlorophenol (2,4-DCP) [48]. The surface
sented the highest phenol uptake capacity; nitric acid treated car- modification of the activated carbon using ammonia was shown
bons and oxygen gasified samples presented an acidic surface to be able to increase its adsorption capacity for 2,4-DCP from
functionality and a low phenol adsorption capacity. The beneficial 232.56 to 285.71 mg/g (22.86% higher) which was explained due
role of nitrogen on phenol adsorption was attributed to adsorbate– to the basic surface functional groups created by nitrogen-incorpo-
adsorbent interactions. ration (via ammonia treatment) which rendered the carbon more
The adsorption of Cr(VI) was examined with NaOH treated AC basic; and created a surface that was more positively charged
[54]. The researchers noted the reduction rates of Cr(VI) by a factor and thus improved the activated carbon uptake.
of about 2 in 170 min and suggested that this reduction might be Chen et al. [59] investigated the modification of AC surface with
due to the decrease of specific surface area or micropore volume. ammonia thermal treatment to achieve higher adsorption capacity

Under alkaline environment (basic components in aqueous solu- for perchlorate (ClO4 ) from aqueous solution, without undermin-
tions), it was expected that OH ion reacted with the surface func- ing the pore structure that was beneficial to the perchlorate
tional groups of AC. Even though NaOH treatment increased the adsorption. They reported that the most favorable improvement
numbers of various surface oxygen complexes, reduction rates of occurred between 650 and 700 °C in which a four-fold increase

Cr(VI) underwent no significant improvement because the specific in perchlorate adsorption was observed. The ClO4 sorption on acti-
surface area decreased and the basic Cr(VI) species in solution vated carbon was strongly correlated with the net positive charge
underwent electron acceptor (acid)-electron donor (base) intermo- on the carbon surface and an increase in the number of cationic
lecular interaction between adsorbate and adsorbent. Chiang et al. sites. However, it has been argued by some workers [60] that acti-
[55] also verified that AC treated with NaOH showed major in- vated carbon contains both cationic, basic groups, most likely pyr-
crease in concentration of phenolic functional groups on the sur- one groups, and anionic, acidic groups, mainly carboxylic groups,
face. Przepiorski [56] studied the influence of AC treatment with and the effect of ammonia or similar basic treatment on the net po-
gaseous ammonia from 400 to 800 °C on phenol adsorption from sitive charge of the carbon surface can also be explained by a de-
aqueous solution. It was found that the adsorption capacity toward crease in the number of anionic groups on the activated carbon
phenol was enhanced as much as 29% at the optimum treatment surface. They further stated that in addition to remove anionic
temperature (700 °C). Those activated carbons containing larger groups from the carbon surface, the chemical modifications might
micropore volumes showed higher ability to adsorb phenol. alter the cationic groups on the carbon surface. Ammonia tailoring
The adsorption of p-chlorophenol (PCP) from aqueous solution introduced quaternary nitrogen sites into the interior of the graph-
on ACs with basic surface properties has been studied [57]. The ene sheets. The surface chemistry of ordered mesoporous carbon
ACs were prepared by two methods. The first method was based CMK-3 was successfully modified by ammonia heat treatment at
on the modification of a commercial CWZ AC by high temperature high temperature of 1173 K [61]. It was reported by the authors
treatment in an atmosphere of ammonia, nitrogen and hydrogen. that four types of basic nitrogen functional groups were created
The second series was produced through carbonization and subse- on the carbon surface, (i) pyridine-like nitrogen and (ii) aromatic
quent activation of N-polymer-based chars using CO2 and steam as amines (being the dominant functional groups), (iii) quaternary
activation agents. Basically, all the ACs were microporous in nat- nitrogen, and (iv) protonated amide by making the carbon ‘‘more
ure, however, CWZ ACs were discriminated by a higher contribu- basic’’. This functionalized adsorbent was found to enhance the
tion of wider micropores and mesopores. The basic surface adsorption capacity of three anionic dyes, namely, orange II, RR2,
properties were supported by a predominant contribution of basic and AB1, by 90–200% and 40–60% compared to the commercial
sites, 0.90–1.56 mmol/g compared to 0.11–0.53 mmol/g of acidic activated carbon and unmodified CMK-3, respectively. This signif-
sites and a relatively high pHPZC, being in the range of 8.2–11.3. icant improvement was mainly attributed to the enhanced disper-
The basicity of heat-treated CWZ ACs was induced mostly by the sive forces between the carbon surface and the dye molecules. Two
oxygen functionalities removal, leading to an increase in electron mechanisms, i.e., electrostatic forces and dispersive interactions,
density, while that of N-polymer-based ACs was related to the high were involved in the adsorption of organic dye molecules on car-
amount of nitrogen (2.42–5.42 wt.%). Remarkably shorter time of bons. The latter emerged from the interactions between the delo-
reaching equilibrium in the PCP adsorption on CWZ ACs than on calized p electrons of the Lewis basic sites in the basal planes of
N-polymer-based ACs proved that the kinetics was controlled by the carbon and the free electrons of the dye molecules present in
the pore size distribution. The highest surface coverage was ob- the aromatic rings and multiple bonds.
tained for the CWZ heat-treated with H2 and N2, which were char- Kasnejad et al. [62] proposed a new procedure for nitrogen-
acterized by the lowest heteroatom content. The treatment ation of commercial activated carbon by applying: heat treatment
resulting in reduced amount of polar sites and enhanced electron of adsorbent under the atmosphere of NH3 after pre-oxidation
density on the carbon surface was therefore recommended as a with HNO3. The influence of this treatment on the characteristics
way to improve PCP adsorption. The results of this study indicated and adsorption properties of activated carbon toward Cu(II) re-
that p–p interactions between the phenol ring and the graphene moval were investigated. Comparison of the prepared sample
layers were mainly responsible for the adsorption on the AC sur- with another one modified by NH3 without pre-treatment,
face from aqueous solution. Nitrogen-derived basic sites were ob- showed that the pre-oxidation of adsorbent had a good influence
served to had a minor contribution to the PCP uptake. on increasing the amount of nitrogen functional groups intro-
Commercial activated carbon and activated carbon fiber were duced onto structure of adsorbent. The higher adsorption rate
modified by high temperature helium or ammonia treatment, or and capacity toward Cu(II) was obtained in the case of nitrogen-
iron impregnation followed by high temperature ammonia treat- ated carbons.
ment [58]. Iron-impregnated and ammonia-treated activated car- Overall, it can be concluded that under alkaline (basic) condi-
bons showed significantly higher dissolved organic matter tions, it is expected that OH ions react with the surface functional
uptakes than the virgin activated carbon. The enhanced dissolved groups of AC and such treatments also create more positive charge
organic matter uptake by iron-impregnated AC was suggested on activated carbon surface which are favorable for enhancing the
due to the presence of iron species on the carbon surface. The high- negatively charged species from water.
504 A. Bhatnagar et al. / Chemical Engineering Journal 219 (2013) 499–511

2.3. Impregnation Fe-impregnated activated carbon (Fe–AC) was prepared by

loading of Fe(III) on activated carbon (AC) at different preparations
The term impregnation can be defined as the fine distribution of of pH and was tested for copper removal [73]. Fe–AC had a greater
chemicals and/or metal particles in the pores of AC. The impregna- content of Fe(III) as the pH decreased and showed above
tion of AC with metals such as silver [63], copper [64], aluminum 10,000 mg/kg below pH 3.0. In the pH-edge adsorption test, as
[65] and iron [66] is gaining wide interest because of their signifi- the copper concentration increased, the removal curves shifted to
cantly high adsorption capacity. Impregnated ACs have shown en- the higher pH region due to the finite number of adsorption sites
hanced adsorption potential towards fluoride, cyanide and heavy on the Fe–AC. Cu(II) adsorption was found higher on Fe–AC than
metals like arsenic in water. Some of the important findings are AC. This trend was explained by the additional removal of Cu(II)
being discussed in the subsequent paragraphs. by impregnated Fe(III) through surface complexation reaction.
Iron salts have been widely used in preparing the impregnation The model predictions were generally well simulated the experi-
of AC. Huang and Vane [67] modified the AC with iron salt solution mental adsorption trends below pH 5.5, but severe deviation was
to improve its arsenic removal capacity and a 10-fold increase was observed between the model prediction and the experimental cop-
achieved by optimal pre-treatment procedure over the untreated per removal above pH 5.5, predicting copper precipitation on the
AC. The enhancement in arsenic removal was attributed to adsorp- Fe–AC surface. A model prediction for Cu(II) adsorption, consider-
tion of ferrous ions and formation of arsenate complexes. In a sep- ing Cu(OH)2 precipitation, agreed well with the experimental cop-
arate study, Dastgheib et al. [68] investigated the impregnation of per removal. From the column test, Fe–AC was regarded as a
AC with iron to enhance the removal of dissolved natural organic promising filter material in the treatment of acidic wastewater
matter (DOM) from natural waters. They reported that iron contaminated with copper with proper controlling of the contact
impregnation followed by high temperature ammonia treatment time between copper and Fe–AC. In another study, iron-impreg-
enhanced the DOM uptake about 50–120% (on a mass basis). Gho- nated activated carbons have been found very effective in arsenic
rishi and Keeney [69] investigated the effect of impregnating AC removal [74]. Surface oxidation of carbon by HNO3/H2SO4 or by
with Cl (hydrogen chloride) on adsorption of elemental mercury HNO3/KMnO4 increased the amount of iron that could be loaded
and reported that the adsorption capacity was found to increase to 7.6–8.0%; arsenic stayed below 10 ppb until 12,000 bed volumes
by factors from 2 to 3 over virgin AC in fixed-bed sorption of mer- during rapid small-scale column tests (RSSCTs) using Rutland, MA
cury species. The removal of arsenic from simulated contaminated groundwater (40–60 mg/L arsenic, and pH of 7.6–8.0). Boehm titra-
groundwater by the adsorption onto Fe3+ impregnated granular tions showed that surface oxidation greatly increased the concen-
activated carbon (GAC–Fe) as well as untreated granular activated tration of carboxylic and phenolic surface groups. Iron
carbon (GAC) in presence of Fe2+, Fe3+, and Mn2+ was studied [70]. impregnation by precipitation or iron salt evaporation was also
The equilibrium time for optimum removal of arsenic species was evaluated. Iron content was increased to 9–17% with internal
noted as 8 h for GAC–Fe and 12 h for GAC. Maximum removal of iron-loading, and to 33.6% with both internal and external iron
As(V) and As(III) was observed in the pH range of 5–7 and 9–11, loading. These iron-tailored carbons reached 25,000–34,000 bed
respectively, for both the adsorbents. Under the experimental con- volumes to 10 mg/L arsenic breakthrough during RSSCTs. With
ditions at 30 °C, the optimum removal of As(III), As(V), Fe, and Mn the 33.6% iron loading, some iron was found to be peeled off.
were 93%, 98%, 100%, and 41%, respectively, when GAC–Fe was Besides iron, some other metals such as Al, Ag, Ni, Mn have
used. Percentage removal of arsenic species on GAC and GAC–Fe also been used for impregnating the AC. Leyva Ramos et al. [75]
decreased with the increase in temperature. Using GAC–Fe, the ar- investigated the efficiency of aluminium-impregnated AC for
senic concentration in the treated water could be reduced below removing fluoride from aqueous solutions which was prepared
10 and 50 mg/L from arsenic solutions containing maximum As va- by impregnation with an aluminium nitrate solution at a fixed
lue of 250 and 520 mg/L, respectively. pH, followed by calcination in an inert environment at tempera-
Chang et al. [71] reported a new multi-step procedure for tures above 300 °C. They found that the impregnated AC had a
impregnation of GAC using ferrous chloride as the precursor for fluoride adsorption capacity of three to five times higher than that
removing arsenic from drinking water. Impregnated iron formed of virgin AC and that the solution pH affected the adsorption of
nano-size particles, and existed in both crystalline (akaganeite) aluminium onto AC and the optimum impregnating pH was found
and amorphous iron forms. Iron-impregnated GACs (Fe–GACs) to be 3.5. At pH below 3.0, the aluminium did not adsorb onto the
were treated with sodium hydroxide to stabilize iron in GAC and AC and at pH above 4.0, impregnation was not possible due to
impregnated iron was found very stable at the common pH range precipitation of aluminium. The efficiency of silver and/or nickel
in water treatments. Synthetic arsenate-contaminated drinking impregnated carbon columns for removing cyanide from aqueous
water was used in isotherm tests to evaluate arsenic adsorption solutions was also evaluated [76]. The adsorption isotherms of
capacities. The iron use efficiency maintained at a high level (40– the impregnating elements (Ag and Ni) from aqueous solutions
46 mg As/g Fe) when iron content was less than 4.22%; it then de- on virgin activated carbon were measured. The amount of
creased to 14 mg As/g Fe when iron content increased to 12.13%. adsorbed silver on plain carbon reached 45.7 mg/g carbon and
Fe–GACs performed better in the acidic condition than the basic for nickel was 4.3 mg/g carbon. These impregnated materials were
condition on arsenate adsorption; specifically, the arsenate re- packed in fixed bed columns and used for cyanide removal from
moval rate maintained close to 100% with pH less than 6.0 and de- aqueous solutions. The adsorption was monitored from the break-
clined quickly at pH above 7.0. Different activated carbons through, which indicated that the carbon was no longer adsorbing
modified with iron hydro(oxide) nanoparticles were also tested effectively. The results indicated that carbon–Ag impregnation had
for their ability to adsorb arsenic from water [72]. The surface cyanide removal capacity of nearly two times that of carbon–Ni
areas of the modified activated carbons ranged from 632 m2/g to impregnation and of four times that of plain activated carbon.
1101 m2/g, and their maximum arsenic adsorption capacity varied The researchers suggested that cyanide was probably eliminated
from 370 lg/g to 1250 lg/g. Temperature did not show any signif- in the forms AgðCNÞ 2 and Ni(CN)4 complexes. However, calcina-
icant effect on arsenic adsorption; however, arsenic adsorption tions of impregnated activated carbon under nitrogen at 300 °C
was found to decrease ca. 32% when the solution pH increased showed similar results to impregnated non-calcinated ones.
from 6.0 to 8.0. In addition, when groundwater was used in the Activated carbon was prepared from an inexpensive and renewable
experiments, the arsenic adsorption considerably decreased due carbon source, Typha orientalis, by H3PO4 activation and then
to the presence of competing anions (mainly SO42, Cl and F). impregnated with different Mn salts and tested for its neutral
 A. Bhatnagar et al. / Chemical Engineering Journal 219 (2013) 499–511 505

red (NR) adsorption capacities [77]. The amount of Mn impreg- rect contact between the microwave heating source and the heated
nated in the activated carbon was influenced by the anion species. material; ease of heating process control; high temperature capa-
The maximum monolayer adsorption capacities were obtained as bilities; time and energy savings; and increase of chemical reactiv-
217.39 mg/g and 285.71 mg/g from TOC-Mn(NO3)2 at 298 K and ity [83]. Moreover, microwave processing systems are also
318 K, respectively. relatively compact, portable, maintainable and cost effective.
Li et al. [78] studied mercury adsorption by cupric chloride Liu et al. [84] prepared modified bamboo-based activated car-
impregnated activated carbon, which used water, acetone, and iso- bon in a microwave oven under N2 atmosphere. A gradual decrease
propyl alcohol as solutions. The results indicated that the adsorp- in the surface acidic groups was observed during the modification,
tion of Hg0 on the cupric chloride impregnated activated carbon while the surface basicity was enhanced to some extent, which
in isopropyl alcohol was the greatest. Accumulation of Cu on the gave rise to an increase in the pHpzc. An increase in the micropores
modified activated carbon surface and other oxygen functional was observed in the beginning, and micropores were then ex-
groups was suggested as the key factor in affecting the chemisorp- tended into larger ones, resulting in an increase in the pore volume
tion of mercury. A thermal hydrolysis method was developed to and average pore size. Adsorption studies showed enhanced
anchor iron hydro(oxide) nanoparticles onto granular activated adsorption of methylene blue on the modified activated carbons,
carbons for the removal of As(V) from water [79]. The authors caused mainly by the enlargement of the micropores. Adsorption
did not find any direct relationship between iron content and ar- isotherm fittings revealed that Langmuir and Freundlich models
senic adsorption capacity. The arsenic uptake was found to be were applicable for the virgin and modified activated carbons,
3.25, 1.45 and 1.65 mg As/g of F400-Fe, ACZ–Fe and ACP–Fe respec- respectively. Kinetic studies exhibited faster adsorption rate of
tively at pH 7.0, 25 °C and 1 ppm of arsenic concentration at equi- methylene blue on the modified activated carbons. The feasibility
librium. Two series of activated carbons modified by Fe(II) and of preparing activated carbon from corncob furfural residue with
Fe(III) were used as adsorbents for the removal of phosphate in ZnCl2 by microwave irradiation was studied [85]. The effect of
aqueous solutions [80]. Maximum removal of phosphate was ob- the ratio by weight of ZnCl2 to corncob furfural residue, ZnCl2 solu-
tained in the pH range of 3.78–6.84 for both adsorbents. Results tion soaking time, microwave irradiation time and the pH value of
suggested that the main phase formed in modified carbons was ZnCl2 solution on the quality of activated carbon was investigated.
goethite (modified by Fe(II)) and akaganeite (modified by Fe(III)), On the condition that the microwave power was 800 W and ZnCl2
and the presence of iron oxides significantly affected the surface solution mass concentration was 50%, the best technological
area and the pore structure of the activated carbon. AC modified parameter on preparing activated carbon with ZnCl2 by microwave
by Fe(II) showed higher phosphate removal capacity than AC mod- irradiation was obtained: at the ratio of 3.5:1 by weight of ZnCl2 to
ified by Fe(III), which could be attributed to its better intra-particle corncob furfural residue, microwave irradiation time was 20 min,
diffusion and higher binding energy. ZnCl2 solution soaking time was 12 h, the pH value of ZnCl2 solu-
The activated carbons were prepared using steam activation tion was 2.0, the activated carbon yield reached 33.1% and the
method in a pilot scale set-up for the bio-char obtained from fast decolorizing capacity for methylene blue was 202.5 mg/g, the
pyrolysis of whitewood and obtained ACs were impregnated using product was used in treating Cr6+ solution (pH value 4.0, concen-
various precursors such as potassium and ammonium halides, with tration 50 mg/L), its adsorption capacity of Cr6+ was 7.583 mg/g.
the halide loadings in the range of 0.1–1.0 mol% of carbon content Various other researchers also modified the surface chemistry of
[81]. It was observed that impregnation of the halide ions signifi- activated carbons by means of microwave-induced treatments
cantly improved the performance of the AC for elemental mercury [86,87]. Considerable research has also been made by Foo and
capture in nitrogen flow. For the same molar loading of halide ions, Hameed to prepare microwave assisted ACs for the removal of var-
mercury removal efficiency increased in the following order: ious pollutants from water and wastewater [88–97].
AC < Cl-impregnated AC < Br-impregnated AC < I-impregnated AC.
Also, Hg removal efficiency of ammonium halide impregnated acti- 2.5. Ozone treatment
vated carbon was higher compared to potassium halide impreg-
nated activated carbon, which was due to the better access of Hg One of the most widely used oxidants is ozone for the depura-
to active sites in pores for ammonium halide impregnating agent. tion of toxic organic compounds present in water. The simulta-
Changes in Hg removal efficiency with the process temperature neous use of ozone and AC has recently been proposed as an
showed that possibly the main mechanism for Hg capture by acti- alternative to ozonation [98]. Initially the high adsorption capacity
vated carbon changed from physisorption to chemisorption after of AC was considered responsible for the higher efficacy of the
impregnation. combined system. However, some researchers [99,100] reported
It can be concluded that research conducted on impregnated that other phenomena besides adsorption increase the efficiency
ACs has shown significant results and the main advantages of of the combined system, such as: (i) reaction between the ozone
impregnating the AC include the optimization of the catalytic and the organic matter adsorbed on the AC and, (ii) the generation
properties of AC by promoting its built-in catalytic oxidation capa- of free radicals from the reaction between the dissolved ozone and
bility and to promote synergism between AC and the impregnating the AC, which can produce mineralization of the organic matter.
agent. However, more research in the area of leaching of impreg- Rivera-Utrilla and Sanchez-Polo [101] assessed the adsorption
nated metals on AC should be conducted to ensure the feasibility of 1-naphthalenesulphonic (NS), 1,5-naphthalenedisulphonic
of the process. (NDS) and 1,3,6-naphthalenetrisulphonic (NTS) acids on ozonated
activated carbons. Commercial activated carbon (Filtrasorb 400)
2.4. Microwave treatment was treated with different ozone doses in order to study the effect
of ozone treatment on their surface properties, and to investigate
In recent years, modification of AC by means of microwave radi- the behavior of these carbon samples in the adsorption of the
ation is gaining wide attention due to its capacity in heating at naphthalenesulphonic acids. Experimental results revealed that
molecular level leading to homogenous and quick thermal reac- the adsorption capacity sharply decreased as the number of sul-
tions [82]. Compared with conventional heating, microwave heat- phonic groups in the aromatic ring increased. As the concentration
ing offers many advantages such as: microwave energy heats the of oxygenated electron-withdrawing groups on the carbon surface
material from inside out; there is no need for heat convection increased, a significant reduction in adsorption capacity of aro-
through a fluid; microwave energy provides rapid heating; no di- matic sulphonic compounds was observed. These results indicated
506 A. Bhatnagar et al. / Chemical Engineering Journal 219 (2013) 499–511

that the adsorption process took place mainly by p–p dispersion 2.6. Plasma treatment
interactions between the aromatic ring electrons of the naphtha-
lenesulphonic acids and the basal plane of the AC. Moreover, in Plasma oxidation is a process wherein AC is exposed to plasma
all cases, the adsorption capacity of aromatic sulphonic acids de- under vacuum or atmospheric pressure in the presence of con-
creased as the pH increased. This indicated that electrostatic inter- trolled air or oxygen. During this process, very little textural
actions between the adsorbate and the carbon surface also took change takes place; however, there would be a substantial change
place, as a result of the ionisation of oxygenated surface groups in the surface chemistry of AC. During plasma oxidation, there
on the carbon surface. Nevertheless, the dispersive (non-electro- would be an increase in the surface acidity due to chemical addi-
static) forces governed the adsorption processes. The impact of tion of oxygen to carbon surfaces as oxygen free radicals react
ozonation on textural and chemical surface characteristics of two aggressively with carbon atoms located at the peripheral surface
granular activated carbons (GAC), namely F400 and AQ40, and of the graphene layers.
their ability to adsorb phenol (P), p-nitrophenol (PNP), and p-chlo- Granular activated carbon has been treated by the dielectric
rophenol (PCP) from aqueous solutions have been studied [102]. barrier discharge (DBD) plasma for enhancing its adsorbability
The porous structure of the ozone-treated carbons remained prac- for metal ions [105]. A mixture of helium and oxygen was used
tically unchanged with regard to the virgin GAC. However, impor- as a feeding gas to create oxygen radicals in helium plasma. A
tant modifications of the chemical surface and hydrophobicity mechanism of the metal ion adsorption on the AC surface was
were measured from FTIR spectroscopy, pH titrations, and deter- ion exchange between metal cations in the aqueous solution and
mination of pHPZC. The ozone treatment at either room tempera- hydrogen ions on the activated carbon surface. Therefore, the AC
ture (i.e., about 25 °C) or 100 °C gave rise to acidic surface surface was oxidized by the helium–oxygen DBD plasma to in-
oxygen groups (SOG). At 25 °C, primarily carboxylic groups were crease the hydrogen ions mostly released from functional groups
formed while a more homogeneous distribution of carboxylic, lac- on the surface. Oxygen species produced during the discharge react
tonic, hydroxyl, and carbonyl groups was obtained at 100 °C. The on the surface resulting in the creation of weakly acidic functional
experimental isotherms for phenolic compounds on both GAC groups that played an important role in adsorbing metal cations.
were analyzed using the Langmuir model. Dispersive interactions Iron cations (Fe2+) in the acidic solution were used for the adsorp-
between p electrons of the ring of the aromatics and those of the tion experiment, and adsorbed by both untreated and treated ACs.
carbon basal planes were thought to be the primary forces respon- The experimental results showed that adsorbability of the plasma-
sible for the physical adsorption whereas oxidative coupling of treated AC was about 3.8 times higher than that of untreated sam-
phenolic compounds catalyzed by basic SOG was a major cause ple, when the treatment was conducted for 30 min in the helium
of irreversible adsorption. The exposure of both GAC to ozone at DBD with 4% content of oxygen. It was observed through the mate-
room temperature decreased their ability to adsorb P, PNP, and rial analyses (FTIR, XPS, and BET) that the improvement of adsorb-
PCP. However, when ozone was applied at 100 °C, adsorption ability was due to the change in the surface chemical structure of
was not prevented but in some cases (P and PNP on F400), the the AC rather than the modification of the surface physical
adsorption process was even enhanced. structure.
Three different surface treatment methods were used to pro- The oxygen plasma treatment of activated carbon fibers (ACFs)
duce surface oxygen groups on carbon using ozone, air, and was carried out to introduce oxygen-containing groups onto car-
hydrogen peroxide to investigate the mercury adsorption at bon surfaces [106]. It was found that the oxygen functional groups,
varying temperature conditions [103]. It was noticed that ozone such as C6H5OH and O  C@O, were increased after the plasma
treatment dramatically increased the mercury sorption capacity treatment. However, a slight decrease of specific surface area was
by factors up to 134, but the activity was easily destroyed by also noticed. HCl removal efficiency of all ACFs modified by plasma
exposure the carbon surface to water vapor atmosphere or by treatment was higher than that of untreated samples. Authors con-
mild heating. Freshly ozone-treated carbon surfaces were shown cluded that when the plasma treatment was accomplished in the
to oxidize iodide to iodine in solution and this ability faded with range of high oxygen content and well-developed micropores, this
aging. FTIR analysis showed broad CAO stretch features from technique could be useful and economic method of removing
950 to 1300 cm1, which decay upon atmospheric exposure highly toxic hydrochloride gases on carbon surfaces. Besides these,
and were similar to the CAOAC asymmetric stretch features of some other researchers also studied the plasma treatment on AC
ethylene secondary ozonide. The combined results suggested [107,108].
that the ultra-high mercury adsorption efficiency was due to a Application of low temperature oxygen plasma was proposed
subset of labile CAO functional groups with residual oxidizing by Zhang et al. [109] to modify activated carbons for enhancing
power that were likely epoxides or (epoxide-containing) second- their adsorption toward dibenzothiophene (DBT) in a model die-
ary ozonides. sel fuel. After oxidation of AC with low temperature oxygen plas-
The control of the surface chemistry of AC by ozone and heat ma, its surface concentration of the acidic oxygen-containing
treatment was also investigated [104]. ACs were prepared using groups increased greatly. As a result, the adsorption capacities
cherry stones by carbonization at 900 °C and activating in CO2 or of the modified ACs for DBT were greatly improved. Fixed-bed
steam at 850 °C. The obtained products were ozone-treated at breakthrough experiments showed that the adsorption capacity
room temperature. After their thermo gravimetric analysis, the of the modified ACs for DBT increased by 49.1% compared to
samples were heat-treated to 300, 500, 700 or 900 °C. The textural the original AC. Activated carbon fiber (ACF) was modified by a
characterization was carried out by N2 adsorption at 77 K, mercury novel modification technology, gilding arc discharge, while its
porosimetry, and density measurements. The surface analysis was adsorption capacity was studied with the acid orange II (AO II)
performed by the Boehm method and pH of the point of zero solution selected as the target wastewater [110]. The results
charge (pHpzc). It was found that the treatment of AC with ozone showed that the specific surface area and pore volume of ACF de-
combined with heat treatment enabled to control the acidic-basic creased after the plasma treatment, while the amounts of oxy-
character and strength of the carbon surface. Surface charge of gen-containing functional groups on the surface of ACF
these carbons ranged between 3.6 and 10.3. Overall, it can be said increased compared with the raw ACF. Moreover, it was observed
that ozone treatment of AC provides an efficient way of water that the adsorption capacity of the modified ACF was improved
treatment however, more research is needed to make the process by nearly 20.9%. Many researchers opine that the plasma treat-
economically feasible. ment process requires a one-time investment for equipment
 A. Bhatnagar et al. / Chemical Engineering Journal 219 (2013) 499–511 507

and instrumentation and recurring expenses for the electricity re- The performance of bio-polymer modified activated carbon and
quired to generate plasma. commercial activated carbon for palladium and platinum removal
from dilute aqueous solutions was investigated [123]. Commercial
2.7. Biological modification activated carbon showed the palladium and platinum removal po-
tential lower (adsorption capacity of 35.7 and 45.5 mg/g, respec-
The use of biologically modified AC for the removal of aquatic tively) than the bio-polymer modified activated carbon which
pollutants has been the subject of research for the past several dec- was able to adsorb 43.5 and 52.6 mg/g of palladium and platinum,
ades. In this process, bacteria get trapped within the activated car- respectively. Cationic surfactant (cetyltrimethylammonium chlo-
bon and these trapped bacteria begin to multiply in the activated ride (CTAC)) was employed to modify granular activated carbon
carbon chamber under an ideal environment of temperature and (GAC) for bromate (BrO 3 ) removal [124]. Cationic surfactant mod-
organic nutrients for growth. One of the beneficial effects in the ification was proven to be effective in enhancing the adsorption of
biological AC process is that carbon-bed life can be prolonged by bromate. Specifically, virgin GAC could only adsorb about 5.53 mg
converting a portion of recalcitrant organics to biodegradable bromate/g GAC. In contrast, at best, GAC–CTAC-2 (modified by
organics by preozonation. The attached microorganisms then con- 2 mmol/L of CTAC) showed an adsorption capacity of 30.30 mg
vert the biodegradable portion to biomass, carbon dioxide and bromate/g at pH 6.0. In addition, there was an optimum CTAC load-
waste products before this material can occupy adsorption sites ing of about 0.18 mmol/g for the GAC studied. In another study,
on the AC [111]. Biological activated carbon (BAC) has been widely surfactant-modified powdered activated carbon (SM-PAC) was
used for the removal of wide variety of toxic aquatic pollutants e.g., synthesized and used for the removal of bromate from water
17b-estradiol (E2) [112], methyl tert-butyl ether (MTBE) [113], [125]. Three cationic surfactants, cetylpyridinium chloride (CPC),
phenols [114], natural organic matter and organic micropollutants hexadecyltrimethylammonium chloride (CTAC), and hexadecyl-
[115], pesticides [116], bromate [117], etc. trimethyl-ammonium bromide (CTAB) were used for modification.
In particular, CPC- and CTAC-modified SM-PAC exhibited signifi-
2.8. Miscellaneous modification methods of AC cant increases in zeta potential. The surface area of SM-PAC de-
creased after the adsorption of the surfactant at the pores of the
Besides the above mentioned methods, AC has also been modi- PAC but it was not significant. Compared to the virgin PAC, CPC-,
fied by using different other chemicals. Some of these methods are and CTAC-modified SM-PAC exhibited high bromate adsorption
being discussed in subsequent paragraphs. The feasibility of cat- capacity and fast bromate adsorption rate. BrO 3 adsorption on
ionic surfactant-modified activated carbon for the removal of SM-PAC and PAC was strongly affected by pH. BrO 3 showed a
Cr(VI) from aqueous solution was studied by Choi et al. [118]. It strong affinity for the surfaces of SM-PAC and virgin PAC at pH be-
was reported by the researchers that the AC modified by hexade- low 7, but BrO 
3 removal decreased as pH increased. BrO3 was re-
cyltrimethylammonium had a higher adsorption capacity for Cr(VI) moved by SM-PAC through electrostatic interaction and ion
than that modified by cetylpyridinium. It was suggested by the exchange reaction. CPC and CTAC modified PAC removed BrO 3
authors that cationic surfactant-modified activated carbon could mainly by ion exchange, while CTAB-modified PAC removed BrO 3
be used to remove anionic Cr(VI) because there were greater num- by both ion exchange and electrostatic interaction.
ber of positively charged adsorption sites compared to AC. 1-acyl- Recently, a novel method has been developed for anchoring
thiosemicarbazide-modified activated carbon (AC–ATSC) was quaternary ammonium/epoxideforming compounds (QAE) within
prepared as a solid-phase extractant and applied for the removal bituminous and wood based GACs for perchlorate removal from
of trace Cu(II), Hg(II) and Pb(II) by Gao et al. [119]. The maximum perchlorate-spiked natural groundwater [126]. By means of mul-
static adsorption capacity of Cu(II), Hg(II) and Pb(II) onto the AC– tiple analyses, the authors surmised that the active quaternary
ATSC were 78.20, 67.80 and 48.56 mg/g, respectively at pH 3. The ammonium sites within the tailored bituminous-based GAC were
adsorbed metal ions were quantitatively eluted by 3.0 mL of 2% more exposed and thus more accessible for perchlorate adsorp-
CS(NH2)2 and 2.0 mol/L HCl solution. It was reported that common tion than those within wood based GAC. The performance of a
coexisting ions did not interfere with the separation. The removal conventional (F400) and a surface modified activated carbon
of heavy metals from industrial phosphoric acid was examined by (F400AN) has been examined for the sorption of benazolin and
using activated carbons modified with sodium dodecylsulphate 2,4-dichlorophenoxy acetic acid (2,4-D) from water [127]. It
(SDS) and with sodium diethyl dithiocarbamate (SDDC) in fixed was observed that the modified carbon, F400AN, which was ob-
bed [120]. Fixed bed activated carbon columns modified with tet- tained by annealing the conventional sample, had a higher BET
rabutyl ammonium iodide (TBAI) and SDDC were investigated for surface area (960 m2/g compared to 790 m2/g) and it had a high-
the removal of copper, zinc, chromium and cyanide from metal fin- er proportion of micropores. This was attributed to the loss of
ishing wastewater. Tetrabutyl ammonium (TBA)-carbon column oxygen containing functional groups during the thermal treat-
was used for the cyanide removal, while SDDC-carbon column ment. Zeta potential and pH titration measurements also showed
was used for Cu, Zn and Cr removal [11]. A total CN removal that acidic functionality had been lost on the F400AN sample.
was observed when using the TBA-carbon column with uptake The same researchers also extended their study for the adsorp-
capacity of 29.2 mg/g. The uptake capacity of Cu, Zn and Cr metal tion of atrazine from water using a conventional activated car-
ions were found to be 38, 9.9 and 6.84 mg/g, respectively, using bon, F400, an annealed carbon sample, F400AN, and an
SDDC-carbon column. Activated carbon modified with aminated carbon sample, F400NH2 [128]. Characterization of
tris(hydroxymethyl)aminomethane (AC–TRIS) was used for selec- the carbon samples showed that sample F400NH2 had the high-
tive separation of Au(III) [121]. The optimum pH for the separation est proportion of micropores, but had the lowest values of point
of Au(III) by the adsorbent was found to be 1.0, and the maximum of zero charge (PZC) and iso-electric point (IEP). This was attrib-
static adsorption capacity of Au(III) onto the AC–TRIS was uted to the existence of a high proportion of oxygen containing
33.57 mg/g at this pH and 1 h contact time. Zincon-modified acti- functional groups. Adsorption data showed that sample F400AN
vated carbon (AC–ZCN) as a solid-phase was used for simultaneous was superior in the adsorption of atrazine to samples F400 and
preconcentration of trace Cr(III) and Pb(II) [122]. At pH 4, the max- F400NH2. Authors pointed out that the adsorption of atrazine
imum adsorption capacity of Cr(III) and Pb(II) were 17.9 and on activated carbon was not only a function of pore size distribu-
26.7 mg/g, respectively, onto the AC–ZCN. The adsorbed metal ions tion of the adsorbents but also surface functionality had an influ-
were quantitatively eluted by 1 mL of 0.1 mol/L HCl. ence on the performance of the adsorbents.
508 A. Bhatnagar et al. / Chemical Engineering Journal 219 (2013) 499–511

Table 1
List of main methods used to prepare surface modified ACs for the removal of different aquatic pollutants.

Treatment step Main Results Reference


Modified with tetrabutyl ammonium iodide (TBAI) and sodium diethyl A total CN removal was achieved using the TBA-carbon column with [11]
dithiocarbamate (SDDC) uptake capacity of 29.2 mg/g. Using SDDC-carbon column, Cu, Zn and Cr
metal ions were removed with uptake capacity of 38, 9.9 and 6.84 mg/g,
respectively
Oxidized with nitric acid, ammonium persulphate and hydrogen peroxide in Adsorption of Cr(III) increased on oxidation and decreased on degassing, [29]
the solution phase and with gaseous oxygen at 350 °C. The activated while the adsorption of Cr(VI) decreased on oxidation and increased on
carbons were also degassed at different temperatures between 400 °C degassing
and 950 °C
Oxidized with a saturated solution of (NH4)2S2O8 in 1 M H2SO4 Decrease in surface area and carbon porosity after oxidation, but Zn(II) [30]
uptake greatly increased
Oxidation with HNO3 Carboxylic acid groups were the major surface species incorporated and [31]
phenol and quinone groups were introduced during oxidation process.
Cadmium adsorption was dramatically enhanced after the carbon was
oxidized
Heat treatment (1200 K) in nitrogen (N2), air oxidation (693 K), and nitric Oxygen surface complexes, possibly lactone and carbonyl groups, were the [32]
acid (6 N) active sites for Hg0 uptake
Heating from ambient temperature to 900 °C in SO2; treatment at ambient The amount of sulphur introduced was high when heated to 900 °C in SO2. [33]
temperature in SO2; or successive treatments in SO2 and H2S at ambient Heating to 900 °C was found to be the most effective treatment, and the
temperature adsorption in this case was 70.3%.
Oxidation with nitric acid Surface area and pore volume were reduced after oxidation treatment. The [35]
modified sample displayed cation-exchange properties over a wide range of
pH values and exhibited polyfunctional nature.
Treated with HNO3 and NaOH Nitric acid oxidization produced positive acid groups, and subsequently [36]
sodium hydroxide treatment replaced H+ of surface acid groups by Na+, and
the acidity of activated carbon decreased. Adsorption capacity of Cr(VI)
increased from 7.61 mg/g to 13.88 mg/g.
Treated with 1.0 M citric acid, followed by an optional step of reaction with Surface modification reduced the specific surface area by 34% and point of [38]
1.0 M NaOH zero charge (pHpzc) of the carbon by 0.5 units but increased the Cu(II)
adsorption capacity to 14.92 mg/g, which was 140% higher than the
unmodified carbon
Ammonia-modification Increased adsorption capacity for 2,4-DCP from 232.56 to 285.71 mg/g [48]
(22.86% higher)
Treatment with gaseous ammonia from 400 to 800 °C Adsorption capacity toward phenol was enhanced by as much as 29% at the [56]
optimum treatment temperature (700 °C)
Modified by high temperature helium or ammonia treatment, or iron Iron-impregnated and ammonia-treated activated carbons showed [58]
impregnation followed by high temperature ammonia treatment significantly higher dissolved organic matter uptakes than the virgin
activated carbon
Impregnation of AC with iron followed by high temperature ammonia DOM uptake enhanced about 50–120% [68]
treatment
Impregnation with an aluminium nitrate solution, followed by calcinations Fluoride adsorption capacity increased of three to five times higher than [75]
in an inert environment at temperatures above 300 °C that of virgin AC
Microwave treatment under N2 atmosphere Enhanced adsorption of methylene blue caused mainly by the enlargement [84]
of the micropores
Ozone treatment Mercury sorption increased [103]
Oxygen plasma treatment Oxygen functional groups were increased after the plasma treatment, [106]
however, a slight decrease of specific surface area was shown
Modified with tris(hydroxymethyl)aminomethane Adsorption capacity of Au(III) onto AC–TRIS was 33.57 mg/g at pH 1.0 [121]

Various researchers also prepared and applied modified forms of methods have been found promising in enhancing the potential of
activated carbons in the application of metal enrichment and water ACs for the removal of target aquatic pollutants and make them ver-
treatment. Activated carbons functionalized with pyrocatechol satile in water pollution control.
violet [129], 1,2-cyclohexanediondioxime [130], bis-salicyl-alde-
hyde-1,3-propan-diimine (BSPDI) [131], tartrazine [132], 1,10-phe-
nanthroline [133], dithioxamide [134], and N,N-etylenebis-(ethane 3. Cost estimation
sulfonamide) [135] were used as a chelating collector for various
metal ions. Also, different complexing agents, such as, ammonium Adsorbent’s cost is one of the crucial and important factors to
pyrrolidine dithiocarbamate [136], and 5,5-diphenylimidazoli- know the economic feasibility of the process and designing and
dine-2,4-dione (phenytoin) [137] have been used as modifying installing water and wastewater treatment plants; however, it is
agents for AC for the removal of different metal ions. In addition, seldom reported in the research papers. Especially no information
several heavy metal ions enrichment has been achieved after mod- about the cost has been reported in the literature for surface mod-
ification of AC with chelating agents, such as pyrocatechol violet for ified activated carbons. It has been pointed out earlier by Gupta
Cu, Mn, Co, Cd, Pb, Ni, and Cr [129], 8-hydroxyquinoline for Cd [138], et al. [142] that several factors are responsible for the cost of the
Co, Hg, and Ni [139], pyridyl azo resorcinol for Cu, Co, Cd, Cr, Ni, Pb, precursor and/or the final material (sorbent) which include its
and, V [140], Tris(hydroxymethyl)aminomethane for Cr(VI) [141]. availability; form (natural, agro-industrial by-products, or synthe-
Overall it can be concluded that research conducted on modifying sized products); the processing required; the treatment condi-
ACs using different chemicals has shown significant results how- tions; and both recycling and lifetime issues [142]. Besides it, the
ever, more research in the area of leaching of chemicals from AC cost also depends on some other factors such as, production period
should be conducted to ensure the feasibility of the process. Table 1 and location (when and where the adsorbents are made), and their
lists some of the main methods used to prepare surface modified application (if they will be used in developed countries, developing
ACs for the removal of different aquatic pollutants. These treatment countries, or underdeveloped countries).
 A. Bhatnagar et al. / Chemical Engineering Journal 219 (2013) 499–511 509

4. Concluding remarks [2] J.R. Perrich, Activated Carbon Adsorption for Wastewater Treatment, CRC
Press, Boca Raton, FL, 1981.
[3] H. Marsh, F.R. Reinoso, Activated Carbon, Elsevier Science, Amsterdam, 2005.
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which have been studied to increase the adsorption potential of Group, New York, NY, 2005.
[5] T.J. Bandosz, Activated Carbon Surfaces in Environmental Remediation
AC for diverse type of pollutants from water and wastewater. These
(Interface Science and Technology), Academic Press, London, UK, 2006.
methods have been found to alter the pore structure of carbon in [6] H. Sontheimer, J. Crittenden, R.S. Summers, Activated Carbon for Water
little or large extent. Generally, acidic treatment of AC has been Treatment, 2nd ed., Forschungstelle Engler – Bunte-Institute, Universität
Karlsruhe, Karlsruhe, Germany, 1988.
found promising for the higher uptake of metal ions, while base
[7] A. Dabrowski, Adsorption-from theory to practice, Adv. Colloid Interface Sci.
treatment of AC has shown higher uptake of anionic species from 93 (2001) 135–224.
aqueous solutions. However, some of the shortcomings of modifi- [8] P.J. Purcell, Milestones in the development of municipal water treatment
cation methods include the associated costs involve in the process, science and technology in the 19th and early 20th centuries: part I, Water
Environ. J. 19 (2006) 230–237.
and leaching of the chemicals used in the modification process in [9] J.W. Hassler, in: P.N. Cheremisinoff, E. F (Eds.), Carbon Adsorption Handbook,
the water being treated. Furthermore, very little information is Ann Arbor Science, Ann Arbor, MI, 1980.
available where these surface modified ACs have been employed [10] R.T. Yang, Adsorbents: Fundamentals and Applications, John Wiley & Sons
Inc., Hoboken, NJ, 2003.
on column, pilot or full scale and most of these studies are limited [11] L. Monser, N. Adhoum, Modified activated carbon for the removal of copper,
to lab-scale only (batch tests). Therefore, assessment of the perfor- zinc, chromium and cyanide from wastewater, Separ. Purif. Technol. 26
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facility, Appl. Environ. Microbiol. 46 (1983) 406–416. [137] M. Ghaedi, F. Ahmadi, Z. Tavakoli, M. Montazerozohori, A. Khanmohammadi,
[112] Z. Li, B. Dvorak, X. Li, Removing 17b-estradiol from drinking water in a M. Soylak, Three modified activated carbons by different ligands for the solid
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dichlorophenol in biological activated carbon system, Environ. Technol. 21 [139] J. Shiowatana, K. Benyatianb, A. Siripinyanond, Determination of Cd, Co., Hg,
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 Biochar Application to Soils
A Critical Scientific Review
of Effects on Soil Properties, Processes and Functions

F. Verheijen, S. Jeffery, A.C. Bastos, M. van der Velde, I. Diafas

EUR 24099 EN - 2010

The mission of the JRC-IES is to provide scientific-technical support to
the European Union’s policies for the protection and sustainable
development of the European and global environment.

European Commission,
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Institute for Environment and Sustainability

Contact information
Address: Dr. Frank Verheijen, European Commission, Joint Research
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JRC 55799

EUR 24099 - EN
ISBN 978-92-79-14293-2
ISSN 1018-5593
DOI 10.2788/472

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Title page artwork: Charcoal drawing by Marshall Short

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 Biochar Application to Soils
A Critical Scientific Review
of Effects on Soil Properties, Processes and Functions

F. Verheijen1, S. Jeffery1, A.C. Bastos2, M. van der Velde1, I. Diafas1

1
Institute for Environment and Sustainability, Joint Research Centre (Ispra)
2
Cranfield University (UK)
*
Corresponding author: frank.verheijen@jrc.ec.europa.eu
ACKNOWLEDGEMENTS
The preparation of this report was an institutional initiative. We have received
good support from Luca Montanarella, our soil colleagues in DG ENV
provided helpful reviews and comments along the way, and two external
experts reviewed the document in detail, thereby improving the quality of the
final version.

This volume should be referenced as: Verheijen, F.G.A., Jeffery, S., Bastos,
A.C., van der Velde, M., and Diafas, I. (2009). Biochar Application to Soils - A
Critical Scientific Review of Effects on Soil Properties, Processes and
Functions. EUR 24099 EN, Office for the Official Publications of the European
Communities, Luxembourg, 149pp.

4
EXECUTIVE SUMMARY
Biochar application to soils is being considered as a means to sequester
carbon (C) while concurrently improving soil functions. The main focus of this
report is providing a critical scientific review of the current state of knowledge
regarding the effects of biochar application to soils on soil properties,
processes and functions. Wider issues, including atmospheric emissions and
occupational health and safety associated to biochar production and handling,
are put into context. The aim of this review is to provide a sound scientific
basis for policy development, to identify gaps in current knowledge, and to
recommend further research relating to biochar application to soils. See Table
1 for an overview of the key findings from this report. Biochar research is in its
relative infancy and as such substantially more data are required before
robust predictions can be made regarding the effects of biochar application to
soils, across a range of soil, climatic and land management factors.

Definition
In this report, biochar is defined as: “charcoal (biomass that has been
pyrolysed in a zero or low oxygen environment) for which, owing to its
inherent properties, scientific consensus exists that application to soil at a
specific site is expected to sustainably sequester carbon and concurrently
improve soil functions (under current and future management), while avoiding
short- and long-term detrimental effects to the wider environment as well as
human and animal health." Biochar as a material is defined as: "charcoal for
application to soils". It should be noted that the term 'biochar' is generally
associated with other co-produced end products of pyrolysis such as 'syngas'.
However, these are not usually applied to soil and as such are only discussed
in brief in the report.

Biochar properties
Biochar is an organic material produced via the pyrolysis of C-based
feedstocks (biomass) and is best described as a ‘soil conditioner’. Despite
many different materials having been proposed as biomass feedstock for
biochar (including wood, crop residues and manures), the suitability of each
feedstock for such an application is dependent on a number of chemical,
physical, environmental, as well as economic and logistical factors. Evidence
suggests that components of the carbon in biochar are highly recalcitrant in
soils, with reported residence times for wood biochar being in the range of
100s to 1,000s of years, i.e. approximately 10-1,000 times longer than
residence times of most soil organic matter (SOM). Therefore, biochar
addition to soil can provide a potential sink for C. It is important to note,
however, that there is a paucity of data concerning biochar produced from
feedstocks other than wood. Owing to the current interest in climate change
mitigation, and the irreversibility of biochar application to soil, an effective
evaluation of biochar stability in the environment and its effects on soil
processes and functioning is paramount. The current state of knowledge
concerning these factors is discussed throughout this report.

Pyrolysis conditions and feedstock characteristics largely control the physico-

chemical properties (e.g. composition, particle and pore size distribution) of

5
the resulting biochar, which in turn, determine the suitability for a given
application, as well as define its behaviour, transport and fate in the
environment. Reported biochar properties are highly heterogeneous, both
within individual biochar particles but mainly between biochar originating from
different feedstocks and/or produced under different pyrolysis conditions. For
example, biochar properties have been reported with cation exchange
capacities (CECs) from negligible to approximately 40 cmolc g-1, C:N ratios
from 7 to 500 (or more). The pH is typically neutral to basic and as such
relatively constant. While such heterogeneity leads to difficulties in identifying
the underlying mechanisms behind reported effects in the scientific literature,
it also provides a possible opportunity to engineer biochar with properties that
are best suited to a particular site (depending on soil type, hydrology, climate,
land use, soil contaminants, etc.).

Effects on soils
Biochar characteristics (e.g. chemical composition, surface chemistry, particle
and pore size distribution), as well as physical and chemical stabilisation
mechanisms of biochar in soils, determine the effects of biochar on soil
functions. However, the relative contribution of each of these factors has been
assessed poorly, particularly under the influence of different climatic and soil
conditions, as well as soil management and land use. Reported biochar loss
from soils may be explained to a certain degree by abiotic and biological
degradation and translocation within the soil profile and into water systems.
Nevertheless, such mechanisms have been quantified scarcely and remain
poorly understood, partly due to the limited amount of long-term studies, and
partly due to the lack of standardised methods for simulating biochar aging
and long-term environmental monitoring. A sound understanding of the
contribution that biochar can make as a tool to improve soil properties,
processes and functioning, or at least avoiding negative effects, largely relies
on knowing the extent and full implications of the biochar interactions and
changes over time within the soil system.

Extrapolation of reported results must be done with caution, especially when

considering the relatively small number of studies reported in the primary
literature, combined with the small range of climatic, crop and soil types
investigated when compared to possible instigation of biochar application to
soils on a national or European scale. To try and bridge the gap between
small scale, controlled experiments and large scale implementation of biochar
application to a range of soil types across a range of different climates
(although chiefly tropical), a statistical meta-analysis was undertaken. A full
search of the scientific literature led to a compilation of studies used for a
meta-analysis of the effects of biochar application to soils and plant
productivity. Results showed a small overall, but statistically significant,
positive effect of biochar application to soils on plant productivity in the
majority of cases. The greatest positive effects were seen on acidic free-
draining soils with other soil types, specifically calcarosols showing no
significant effect (either positive or negative). There was also a general trend
for concurrent increases in crop productivity with increases in pH up on
biochar addition to soils. This suggests that one of the main mechanisms
behind the reported positive effects of biochar application to soils on plant

6
productivity may be a liming effect. However, further research is needed to
confirm this hypothesis. There is currently a lack of data concerning the
effects of biochar application to soils on other soil functions. This means that
although these are qualitatively and comprehensively discussed in this report,
a robust meta-analysis on such effects is as of yet not possible. Table 0.1
provides an overview of the key findings - positive, negative, and unknown -
regarding the (potential) effects on soil, including relevant conditions.

Preliminary, but inconclusive, evidence has also been reported concerning a

possible priming effect whereby accelerated decomposition of SOM occurs
upon biochar addition to soil. This has the potential to both harm crop
productivity in the long term due to loss of SOM, as well as releasing more
CO2 into the atmosphere as increased quantities of SOM is respired from the
soil. This is an area which requires urgent further research.

Biochar incorporation into soil is expected to enhance overall sorption

capacity of soils towards anthropogenic organic contaminants (e.g. polycyclic
aromatic hydrocarbons - PAHs, pesticides and herbicides), in a
mechanistically different (and stronger) way than amorphous organic matter.
Whereas this behaviour may greatly mitigate toxicity and transport of common
pollutants in soils through reducing their bioavailability, it might also result in
their localised accumulation, although the extent and implications of this have
not been fully assessed experimentally. The potential of biochar to be a
source of soil contamination needs to be evaluated on a case-by-case basis,
not only with concern to the biochar product itself, but also to soil type and
environmental conditions.

Implications
As highlighted above, before policy can be developed in detail, there is an
urgent need for further experimental research with regard to long-term effects
of biochar application on soil functions, as well as on the behaviour and fate in
different soil types (e.g. disintegration, mobility, recalcitrance), and under
different management practices. The use of representative pilot areas, in
different soil ecoregions, involving biochars produced from a representative
range of feedstocks is vital. Potential research methodologies are discussed
in the report. Future research should also include biochars from non-lignin-
based feedstocks (such as crop residues, manures, sewage and green waste)
and focus on their properties and environmental behaviour and fate as
influenced by soil conditions. It must be stressed that published research is
almost exclusively focused on (sub)tropical regions, and that the available
data often only relate to the first or second year following biochar application.

Preliminary evidence suggests that a tight control on the feedstock materials

and pyrolysis conditions might substantially reduce the emission levels of
atmospheric pollutants (e.g. PAHs, dioxins) and particulate matter associated
to biochar production. While implications to human health remain mostly an
occupational hazard, robust qualitative and quantitative assessment of such
emissions from pyrolysis of traditional biomass feedstock is lacking.

7
Biochar potentially affects many different soil functions and ecosystem
services, and interacts with most of the ‘threats to soil’ outlined by the Soil
Thematic Strategy (COM(2006) 231). It is because of the wide range of
implications from biochar application to soils, combined with the irreversibility
of its application that more interdisciplinary research needs to be undertaken
before policy is implemented. Policy should first be designed with the aim to
invest in fundamental scientific research in biochar application to soil. Once
positive effects on soil have been established robustly for certain biochars at a
specific site (set of environmental conditions), a tiered approach can be
imagined where these combinations of biochar and specific site conditions are
considered for implementation first. A second tier would then consist of other
biochars (from different feedstock and/or pyrolysis conditions) for which more
research is required before site-specific application is considered.

From a climate change mitigation perspective, biochar needs to be

considered in parallel with other mitigation strategies and cannot be seen as
an alternative to reducing emissions of greenhouse gases. From a soil
conservation perspective, biochar may be part of a wider practical package of
established strategies and, if so, needs to be considered in combination with
other techniques.

Table 0.1 Overview of key findings (numbers in parentheses refer to relevant sections)

Description Conditions

Empirical evidence of Biochar analogues (pyrogenic BC and charcoal) are found in

substantial quanities in soils of most parts of the world (1.2-1.4)
charcoal in soils exists (long
term)

The principle of improving Anthrosols can be found in many parts of the world, although
normally of very small spatial extent. Contemplation of Anthrosol
soils has been tried
generation at a vast scale requires more comprehensive, detailed
successfully in the past and careful analysis of effects on soils as well as interactions with
other environmental components before implementation (1.2-1.3
and throughout)

Plant production has been Studies have been reported almost exclusively from tropical regions
with specific environmental conditions, and generally for very limited
found to increase
time periods, i.e. 1-2 yr. Some cases of negative effects on crop
significantly after biochar production have also been reported (3.3).
addition to soils
Positives

Liming effect Most biochars have neutral to basic pH and many field experiments
show an increase in soil pH after biochar application when the initial
pH was low. On alkaline soils this may be an undesirable effect.
Sustained liming effects may require regular applications (3.1.4)

High sorption affinity for Biochar application is likely to improve the overall sorption capacity
of soils towards common anthropogenic organic compounds (e.g.
HOC may enhance the
PAHs, pesticides and herbicides), and therefore influence toxicity,
overall sorption capacity of transport and fate of such contaminants. Enhanced sorption
soils towards these trace capacity of a silt loam for diuron and other anionic and cationic
contaminants herbicides has been observed following incorporation of biochar
from crop residues (3.2.2)

Microbial habitat and Biochar addition to soil has been shown to increase microbial
biomass and microbial activity, as well as microbial efficieny as a
provision of refugia for
measure of CO2 released per unit microbial biomass C. The degree
microbes whereby they are of the response appears to be dependent on nutrient avaialbility in
protected from grazing soils

8
 Increases in mycorrhizal Possibly due to: a) alteration of soil physico-chemical properties; b)
indirect effects on mycorrhizae through effects on other soil
abundace which is linked to
microbes; c) plant–fungus signalling interference and detoxification
observed increases in plant of allelochemicals on biochar; or d) provision of refugia from fungal
productivity grazers (3.2.6)

Increases in earthworm Earthworms have been shown to prefer some soils amended with
biochar than those soils alone. However, this is not true of all
abundance and activity
biochars, particularly at high application rates (3.2.6)

The use of biochar Charcoal in Terra Preta soils is limited to Amazonia and have
received many diverse additions other than charcoal. Pyrogenic BC
analogues for assessing
is found in soils in many parts of the world but are of limited
effects of modern biochars feedstock types and pyrolysis conditions (Chapter 1)
is very limited

Soil loss by erosion Top-dressing biochar to soil is likely to increase erosion of the
biochar particles both by wind (dust) and water. Many other effects
of biochar in soil on erosion can be theorised, but remain untested
at present (4.1)

Soil compaction during Any application carries a risk of soil compaction when performed
under inappropriate conditions. Careful planning and management
application
could prevent this effect (4.6)

Risk of contamination Contaminants (e.g. PAHs, heavy metals, dioxins) that may be
present in biochar may have detrimental effects on soil properties
Negatives

and functions. The ocurrence of such compounds in biochar is likely

to derive from either contaminated feedstocks or the use of
processing conditions that may favour their production. Evidence
suggests that a tight control over the type of feedstock used and
o
lower pyrolysis temperatures (<500 C) may be sufficient to reduce
the potential risk for soil contamination (3.2.4)

Residue removal Removal of crop residues for use as a feedstock for biochar
production can forego incorporation of the crop residue into the soil,
potentially leading to multiple negative effects on soils (3.2.5.5)

Occupational health and fire Health (e.g. dust exposure) and fire hazards associated to the
production, transport, application and storage of biochar need to be
hazards
considered when determining the suitability for biochar application.
In the context of occupational health, tight health and safety
measures need to be put in place in order to reduce such risks.
Some of these measures have already proved adequate (5.2)
-1
Reduction in earthworm High biochar application rates of >67 t ha (produced from poultry
litter) were shown to have a negative effect on earthworm survival
survival rates (limited
rates, possibly due to increases in pH or salt levels (3.2.6)
number of cases)

Empirical evidence is Biochar analogues do not exist for many feedstocks, or for some
modern pyrolysis conditions. Biochar can be produced with a wide
extremely scarce for many
variety of properties and applied to soils with a wide variety of
modern biochars in soils properties. Some short term (1-2 yr) evidence exists, but only for a
under modern arable small set of biochar, environmental and soil management factors
management and almost no data is available on long term effect (1.2-1.4)

C Negativity The carbon storage capacity of biochar is widely hypothesised,

although it is still largely unquantified and depends on many factors
Unknown

(environmental, economic, social) in all parts of the life cycle of

biochar and at the several scales of operation (1.5.2 and Chapter 5)

Effects on N cycle N2O emissions depend on effects of biochar addition on soil

hydrology (water-filled pore volume) and associated microbial
processes. Mechanisms are poorly understood and thresholds
largely unknown (1.5.2)

Biochar Loading Capacity BLC is likely to be crop as well as soil dependent leading to potential
incompatibilities between the irreversibility of biochar once applied
(BLC)
to soil and changing crop demands (1.5.1)

Environmental behaviour The extent and implications of the changes that biochar undergoes
in soil remain largely unknown. Although biochar physical-chemical

9
mobility and fate properties and stabilization mechanisms may explain biochar long
mean residence times in soil, the relative contribution of each factor
for its short- and long-term loss has been sparsely assessed,
particularly when influenced by soil environmental conditions. Also,
biochar loss and mobility through the soil profile and into the water
resources has been scarcely quantified and transport mechanisms
remain poorly understood (3.2.1)

Distribution and availability Very little experimental evidence is available on the short- and long-
term occurrence and bioavailability of such contaminants in biochar
of contaminants (e.g. heavy
and biochar-enriched soil. Full and careful risk assessment in this
metals, PAHs) within context is urgently required, in order to relate the bioavailability and
biochar toxicity of the contaminant to biochar type and 'safe' application
rates, biomass feedstock and pyrolysis conditions, as well as soil
type and environmental conditions (3.2.4)

Effect on soil organic matter Various relevant processes are acknowledged but the way these are
influenced by combinations of soil-climate-management factors
dynamics
remains largely unknown (Section 3.2.5)

Pore size and connectivity Although pore size distribution in biochar may significantly alter key
soil physical properties and processes (e.g. water retention,
aeration, habitat), experimental evidence on this is scarce and the
underlying mechanisms can only be hypothesised at this stage (2.3
and 3.1.3)

Soil water Adding biochar to soil can have direct and indirect effects on soil
water retention, which can be short or long lived, and which can be
retention/availability
negative or positive depending on soil type. Positive effects are
dependent on high applications of biochar. No conclusive evidence
was found to allow the establishment of an unequivocal relation
between soil water retention and biochar application (3.1.2)

Soil compaction Various processes associated with soil compaction are relevant to
biochar application, some reducing others increasing soil
compaction. Experimental research is lacking. The main risk to soil
compaction could probably be reduced by establishing a guide of
good practice regarding biochar application (3.1.1 and 4.6)

Priming effect Some inconclusive evidence of a possible priming effect exists in

the literature, but the evidence is relatively inconclusive and covers
only the short term and a very restricted sample of biochar and soil
types (3.2.5.4)

Effects on soil megafauna Neither the effects of direct contact with biochar containing soils on
the skin and respiratory systems of soil megafanua are known, nor
the effects or ingestion due to eating other soil organisms, such as
earthworms, which are likely to contain biochar in their guts (3.2.6.3)

Hydrophobicity The mechnanisms of soil water repellency are understood poorly in

general. How biochar might influence hydrophobicity remains largely
untested (3.1.2.1)

Enhanced decomposition of It is unknow how much subsequent agricultural management

practices (planting, ploughing, etc.) in an agricultural soil with
biochar due to agricultural
biochar may influence (accelerate) the disintegration of biochar in
management the soil, thereby potentially reducing its carbon storage potential
(3.2.3)

Soil CEC There is good potential that biochar can improve the CEC of soil.
However, the effectiveness and duration of this effect after addition
to soils remain understood poorly (2.5 and 3.1.4)

Soil Albedo That biochar will lower the albedo of the soil surface is fairly well
established, but if and where this will lead to a substantial soil
warming effect is untested (3.1.3)

10
TABLE OF CONTENTS

ACKNOWLEDGEMENTS 4
EXECUTIVE SUMMARY 5
TABLE OF CONTENTS 11
LIST OF FIGURES 15
LIST OF TABLES 19
LIST OF ACRONYMS 21
LIST OF UNITS 23
LIST OF CHEMICAL ELEMENTS AND FORMULAS 25
LIST OF KEY TERMS 27
1. BACKGROUND AND INTRODUCTION 31
1.1 Biochar in the attention 33
1.2 Historical perspective on soil improvement 35
1.3 Different solutions to similar problems 37
1.4 Biochar and pyrogenic black carbon 37
1.5 Carbon sequestration potential 38
1.5.1 Biochar loading capacity 40
1.5.2 Other greenhouse gasses 41
1.6 Pyrolysis 42
1.6.1 The History of Pyrolysis 43
1.6.2 Methods of Pyrolysis 43
1.7 Feedstocks 45
1.8 Application Strategies 49
1.9 Summary 50
2. PHYSICOCHEMICAL PROPERTIES OF BIOCHAR 51
2.1 Structural and Chemical Composition 51
2.1.1 Structural composition 51
2.1.2 Chemical composition and surface chemistry 52
2.2 Particle size distribution 54
2.2.1 Biochar dust 56
2.3 Pore size distribution and connectivity 56
2.4 Thermodynamic stability 58
2.5 CEC and pH 58
2.6 Summary 58
3. EFFECTS ON SOIL PROPERTIES, PROCESSES AND
FUNCTIONS 61
3.1 Properties 61
3.1.1 Soil Structure 61
3.1.1.1 Soil Density 61
3.1.1.2 Soil pore size distribution 63
3.1.2 Water and Nutrient Retention 64
3.1.2.1 Soil water repellency 66
3.1.3 Soil colour, albedo and warming 67
3.1.4 CEC and pH 68

11
 3.2 Soil Processes 69
3.2.1 Environmental behaviour, mobility and fate 69
3.2.2 Sorption of Hydrophobic Organic Compounds (HOCs) 72
3.2.3 Nutrient retention/availability/leaching 76
3.2.4 Contamination 78
3.2.5 Soil Organic Matter (SOM) Dynamics 81
3.2.5.1 Recalcitrance of biochar in soils 81
3.2.5.2 Organomineral interactions 82
3.2.5.3 Accessibility 83
3.2.5.4 Priming effect 83
3.2.5.5 Residue Removal 85
3.2.6 Soil Biology 85
3.2.6.1 Soil microbiota 87
3.2.6.2 Soil meso and macrofauna 89
3.2.6.3 Soil megafauna 90
3.3 Production Function 91
3.3.1 Meta-analysis methods 91
3.3.2 Meta-analysis results 93
3.3.3 Meta-analysis recommendations 98
3.3.4 Other components of crop production function 98
3.4 Summary 98
4. BIOCHAR AND ‘THREATS TO SOIL’ 101
4.1 Soil loss by erosion 101
4.2 Decline in soil organic matter 103
4.3 Soil contamination 103
4.4 Decline in soil biodiversity 105
4.6 Soil compaction 106
4.7 Soil salinisation 106
4.8 Summary 107
5. WIDER ISSUES 109
5.1 Emissions and atmospheric pollution 109
5.2 Occupational health and safety 111
5.3 Monitoring biochar in soil 113
5.4 Economic Considerations 113
5.4.1 Private costs and benefits 113
5.4.2 Social costs and benefits 116
5.5 Is biochar soft geo-engineering? 117
5.6 Summary 118
6. KEY FINDINGS 121
6.1 Summary of Key Findings 121
6.1.1 Background and Introduction 124
6.1.2 Physicochemical properties of Biochar 124
6.1.3 Effects on soil properties, processes and functions 125
6.1.4 Biochar and soil threats 127
6.1.5 Wider issues 127
6.2 Synthesis 128
6.2.1 Irreversibility 128
6.2.2 Quality assessment 128
6.2.3 Scale and life cycle 129

12
 6.2.4 Mitigation/adaptation 129
6.3 Knowledge gaps 131
6.3.1 Safety 131
6.3.2 Soil organic matter dynamics 131
6.3.3 Soil biology 132
6.3.4 Behaviour, mobility and fate 132
6.3.5 Agronomic effects 133
References 135

13
14
LIST OF FIGURES

Figure 1.1 Google TrendsTM result of “biochar”, “Terra Preta” and “black
earth”. The scale is based on the average worldwide traffic of
“biochar” from January 2004 until June 2009 (search
performed on 04/12/2009) 33
Figure 1.2 Google TrendsTM geographical distribution of the search
volume index of “biochar” of the last 12 months from June
2008 to June 2009 (search performed on 16/09/2009). Data is
normalised against the overall search volume by country 34
Figure 1.3 Scientific publications registred in Thompson’s ISI Web of
Science indexed for either biochar or bio-char including those
articles that mention charcoal (search performed on
4/12/2009) 35
Figure 1.4 Distribution of Anthrosols in Amazonia (left; Glaser et al., 2001)
and Europe (middle and right; Toth et al., 2008; Blume and
Leinweber, 2004) 35
Figure 1.5 Comparing tropical with temperate Anthrosols. The left half
shows a profile of a fertile Terra Preta (Anthrosol with
charcoal) created by adding charcoal to the naturally-occurring
nutrient poor Oxisol (far left; photo courtesy of Bruno Glaser).
The right half (far right) is a profile picture of a fertile European
Plaggen Soil (Plaggic Anthrosol; photo courtesy of Erica
Micheli) created by adding peat and manure to the naturally-
occurring nutrient poor sandy soils (Arenosols) of The
Netherlands 36
Figure 1.6 Terms and properties of pyrogenic BC (adopted from Preston
and Schmidt, 2006) 38
Figure 1.7 Diagram of the carbon cycle. The black numbers indicate how
much carbon is stored in various reservoirs, in billions of tons
(GtC = Gigatons of Carbon and figures are circa 2004). The
purple numbers indicate how much carbon moves between
reservoirs each year, i.e. the fluxes. The sediments, as
defined in this diagram, do not include the ~70 million GtC of
carbonate rock and kerogen (NASA, 2008) 39
Figure 1.9 A graph showing the relative proportions of end products after
fast pyrolysis of aspen poplar at a range of temperatures
(adapted from IEA, 2007) 44
Figure 2.1 Putative structure of charcoal (adopted from Bourke et al.,
2007). A model of a microcristalline graphitic structure is
shown on on the left and an aromatic structure containing
oxygen and carbon free radicals on the right 51

15
Figure 3.1 Typical representation of the soil water retention curve as
provided by van Genuchten (1980) and the hypothesized
effect of the addition of biochar to this soil 66
Figure 3.2 The percentage change in crop productivity upon application of
biochar at different rates, from a range of feedstocks along
with varying fertiliserco-amendments. Points represent mean
and bars represent 95% confidence intervals. Numbers next to
bars denote biochar application rates (t ha-1). Numbers in the
two columns on the right show number of total ‘replicates’
upon which the statistical analysis is based (bold) and the
number of ‘experimental treatments’ which have been grouped
for each analysis (italics) 93
Figure 3.3 Percentage change in crop productivity upon application of
biochar at different rates along with varying fertiliserco-
amendments grouped by change in pH caused by biochar
addition to soil. Points represent mean and bars represent
95% confidence intervals. Values next to bars denote change
in pH value. Numbers in the two columns on the right show
number of total ‘replicates’ upon which the statistical analysis
is based (bold) and the number of ‘experimental treatments’
which have been grouped for each analysis (italics) 94
Figure 3.4 The percentage change in crop productivity o upon application
of biochar at different rates along with varying fertiliserco-
amendments to a range of different soils. Points shows mean
and bars so 95% confidence intervals. Numbers in the two
columns on the right show number of total ‘replicates’ upon
which the statistical analysis is based (bold) and the number of
‘experimental treatments’ which have been grouped for each
analysis (italics) 95
Figure 3.5 The percentage change in crop productivity of either the
biomass or the grain upon application of biochar at different
rates along with varying fertiliserco-amendments. Points
shows mean and bars so 95% confidence intervals. Numbers
in the two columns on the right show number of total
‘replicates’ upon which the statistical analysis is based (bold)
and the number of ‘experimental treatments’ which have been
grouped for each analysis (italics) 96
Figure 3.6 The percentage change in crop productivity upon application of
biochar along with a co-amendment of organic fertiliser(o),
inorganic fertiliser(I) or no fertiliser(none). Points shows mean
and bars so 95% confidence intervals. Numbers in the two
columns on the right show number of total ‘replicates’ upon
which the statistical analysis is based (bold) and the number of
‘experimental treatments’ which have been grouped for each
analysis (italics) 97

Figure 5.1 Effect of transportation distance in biochar systems with bioenergy

production using the example of late stover feedstock (high

16
revenue scenario) on net GHG, net energy and net revenue
(adopted from Roberts et al., 2009)

17
LIST OF TABLES

Table 0.1 Overview of key findings

Table 1.1 The mean post-pyrolysis feedstock residues resulting from
different temperatures and residence times (adapted from IEA,
2007) 45
Table 1.2 Summary of key components (by weight) in biochar feedstocks
(adapted from Brown et al., 2009) 46
Table 1.3 Examples of the proportions of nutrients (g kg-1) in feedstocks
(adapted from Chan and Xu, 2009) 47
Table 2.1 Relative proportion range of the four main components of
biochar (weight percentage) as commonly found for a variety
of source materials and pyrolysis conditions (adapted from
Brown, 2009; Antal and Gronli, 2003) 52
Table 2.2 Summary of total elemental composition (C, N, C:N, P, K,
available P and mineral N) and pH ranges and means of
biochars from a variety of feedstocks (wood, green wastes,
crop residues, sewage sludge, litter, nut shells) and pyrolysis
conditions (350-500ºC) used in various studies (adapted from
Chan and Xu, 2009) 53
Table 3.1 Pore size classes in material science vs. soil science 63
Table 6.1 Overview of key findings 121

19
20
LIST OF ACRONYMS

BC Black carbon
CEC Cation Exchange Capacity
DOM Dissolved Organic Matter
HOCs Hydrophobic Organic Compounds
NOM Natural (or Native) Organic Matter
NPs Nanoparticles
OM Organic Matter
PAHs Polycyclic Aromatic Hydrocarbons
PCDD/PCDFs Dioxins and furans
(S)OC (Soil) Organic Carbon
SOM Soil Organic Matter
SWR Soil Water Repellency
VOCs Volatile Organic Compounds

21
LIST OF UNITS

µm Micrometer (= 10-6 m)
Bar 1 bar = 100 kPa = 0.987 atm
Cmolc g-1 Centimol of charge (1 cmol kg-1 = 1 meq 100g-1) per
gram
Gt y-1 Gigatonnes per year
J g-1 K-1 Joule (1J = 1 kg m2 sec-2) per gram per Kelvin
J g-1 K-1 Joule per gram per Kelvin
K Kelvin (1 K = oC + 273,15)
kJ mol-1 Kilojoule (= 103 J) per mole (1 mol ≈ 6.022x1023 atoms
or molecules of the pure substance measured)
Mg ha-1 Megagram (= 106 g) per hectare
nm Nanometer (= 10-9 m)
o
C sec-1 Degrees Celsius per second (rate of temperature
increase)
t ha-1 Tonnes per hectare
v v-1 Volume per volume (e.g. 1 ml per 100 ml)
w w-1 Weight per weight (e.g. 1 g per 100 g)

23
LIST OF CHEMICAL ELEMENTS AND FORMULAS

Al Aluminium
Ar Arsenic
C Carbon
CaCO3 Calcium carbonate
CaO Calcium oxide
CH4 Methane
Cl Chlorine
CO2 Carbon dioxide
Cr Chromium
Cu Copper
H Hydrogen
H2 Hydrogen gas
Hg Mercury
K Potassium
K2O Potassium oxide
Mg Magnesium
N Nitrogen
N2O Nitrous oxide
Na2O Sodium oxide
NH4+ Ammonium (ion)
Ni Nickel
NO3- Nitrate (ion)
O Oxygen
P Phosphorus
Pb Lead
S Sulphur
Si Silicon
SiO2 Silica (silicon dioxide)
Zn Zinc

25
LIST OF KEY TERMS
Accelerated soil Soil erosion, as a result of anthropogenic activity, in excess of
erosion natural soil formation rates causing a deterioration or loss of one
or more soil functions
Activated carbon (noun) Charcoal produced to optimise its reactive surface area
(e.g. by using steam during pyrolysis)
Anthrosol (count noun) A soil that has been modified profoundly through
human activities, such as addition of organic materials or
household wastes, irrigation and cultivation (WRB, 2006)
Biochar i) (Material) charcoal for application to soil
ii) (Concept) “charcoal (biomass that has been pyrolysed
in a zero or low oxygen environment) for which, owing
to its inherent properties, scientific consensus exists
that application to soil at a specific site is expected to
sustainably sequester carbon and concurrently
improve soil functions (under current and future
management), while avoiding short- and long-term
detrimental effects to the wider environment as well as
human and animal health."
Black carbon (noun) All C-rich residues from fire or heat (including from coal,
gas or petrol)
Black Earth (mass noun) Term synonymous with Chernozem used (e.g. in
Australia) to describe self-mulching black clays (SSSA, 2003)
Char (mass noun) 1. Synonym of ‘charcoal’; 2. charred organic matter
as a result of wildfire (Lehmann and Joseph, 2009)
(verb) synonym of the term ‘pyrolyse’
Charcoal (mass noun) charred organic matter
Chernozem (count noun) A black soil rich in organic matter; from the Russian
‘chernij’ meaning ‘black’ and ‘zemlja’ meaning ‘earth’ or ‘land’
(WRB, 2006)
Coal (mass noun) Combustible black or dark brown rock consisting
chiefly of carbonized plant matter, found mainly in underground
seams and used as fuel (OED, 2003)
Combustion (mass noun) chemistry Rapid chemical combination of a
substance with oxygen, involving the production of heat and light
(OED, 2003)
Decline in soil (soil threat) Reduction of forms of life living in the soil (both in
biodiversity terms of quantity and variety) and of related functions, causing a
deterioration or loss of one or more soil functions
Decline in soil (soil threat) A negative imbalance between the build-up of SOM
organic matter and rates of decomposition leading to an overall decline in SOM
contents and/or quality, causing a deterioration or loss of one or
(SOM) more soil functions
Desertification (soil threat) land degradation in arid, semi-arid and dry sub-humid
areas resulting from various factors, including climatic variations
and human activities, causing a deterioration or loss of one or
more soil functions
Dust The finest fraction of biochar, rather than the particulate matter
emitted during pyrolysis. This fraction comprises distinct particle
sizes within the micro- and nano-size range.
Ecosystem The capacity of natural processes and components to provide
functions goods and services that satisfy human needs, directly or indirectly
Feedstock (noun) Biomass that is pyrolysed in order to produce biochar
Landslides The movement of a mass of rock, debris, artificial fill or earth down
a slope, under the force of gravity
Nanoparticle (noun) Any particle with at least one dimension smaller than 100
nm (e.g. fullerenes or fullerene-like structures, crystalline forms of

27
 silica, cristobalite and tridymite)
Organic carbon (noun) biology C that was originally part of an organism;
(chemistry) C that is bound to at least one hydrogen (H) atom
Pyrolysis (mass noun) The thermal degradation of biomass in the absence
of oxygen leading to the production of condensable vapours,
gases and charcoal
Soil (mass noun) The unconsolidated mineral or organic matter on the
surface of the earth that has been subjected to and shows effects
of genetic and environmental factors of: climate (including water
and temperature effects), and macro- and microorganisms,
conditioned by relief, acting on parent material over a period of
time (ENVASSO, 2008).
(count noun) a spatially explicit body of soil, usually differentiated
vertically into layers formed naturally over time, normally one of a
specific soil class (in a specified soil classification system)
surrounded by soils of other classes or other demarcations like
hard rock, a water body or artificial barriers (ENVASSO, 2008)
Soil compaction (soil threat) The densification and distortion of soil by which total
and air-filled porosity are reduced, causing a deterioration or loss
of one or more soil functions
Soil contamination (soil threat) The accumulation of pollutants in soil above a certain
level, causing a deterioration or loss of one or more soil functions.
Soil erosion (soil threat) The wearing away of the land surface by physical
forces such as rainfall, flowing water, wind, ice, temperature
change, gravity or other natural or anthropogenic agents that
abrade, detach and remove soil or geological material from one
point on the earth's surface to be deposited elsewhere. When the
term ‘soil erosion’ is used in the context of it representing a soil
threat it refers to ‘accelerated soil erosion’.
Soil functions A subset of ecosystem functions: those ecosystem functions that
are maintained by soil
Usage:
Most soil function systems include the following:
1) Habitat function
2) Information function
3) Production function
4) Engineering function
5) Regulation function
Soil organic matter (noun) The organic fraction of the soil exclusive of undecayed
plant and animal residues (SSSA, 2001)
Soil salinisation (soil threat) Accumulation of water soluble salts in the soil, causing
a deterioration or loss of one or more soil functions.
Soil sealing (soil threat and key issue) The destruction or covering of soil by
buildings, constructions and layers, or other bodies of artificial
material which may be very slowly permeable to water (e.g.
asphalt, concrete, etc.), causing a deterioration or loss of one or
more soil functions
Soil threats A phenomenon that causes a deterioration or loss of one or more
soil functions.
Usage:
Eight main threats to soil identified by the EC (2002) with the
addition of desertification:
1. Soil erosion
2. Decline in soil organic matter
3. Soil contamination
4. Soil sealing
5. Soil compaction
6. Decline in soil biodiversity
7. Soil salinisation
8. Landslides

28
 9. Desertification
Soil water the reduction of the affinity of soils to water such that they resist
repellency wetting for periods ranging from a few seconds to hours, days or
weeks (King, 1981)
Terra Preta (noun) Colloquial term for a kind of Anthrosol where charcoal (or
biochar) has been applied to soil along with many other materials,
including pottery shards, turtle shells, animal and fish bones, etc.
Originally found in Brazil. From the Portuguese ‘terra’ meaning
‘earth’ and ‘preta’ meaning ‘black’.

29
1. BACKGROUND AND INTRODUCTION
Biochar is commonly defined as charred organic matter, produced with the
intent to deliberately apply to soils to sequester carbon and improve soil
properties (based on: Lehmann and Joseph, 2009). The only difference
between biochar and charcoal is in its utilitarian intention; charcoal is
produced for other reasons (e.g. heating, barbeque, etc.) than biochar. In a
physicochemical sense, biochar and charcoal are essentially the same
material. It could be argued that biochar is a term that is used for other
purposes than scientific, i.e. to re-brand charcoal into something more
attractive-sounding to serve a commercial purpose. However, from a soil
science perspective it is useful to be able to distinguish between any charcoal
material and those charcoal materials where care has been taken to avoid
deleterious effects on soils and to promote beneficial ones. As this report
makes clear, the wide variety of soil groups and associated properties and
processes will require specific charcoal properties for specific soils in order to
meet the intention of biochar application. Considering the need to make this
distinction, a new term is required and since biochar is the most common term
currently used, it was selected for this report. The definition of the concept of
biochar used in this report is:
“charcoal (biomass that has been pyrolysed in a zero or low oxygen
environment) for which, owing to its inherent properties, scientific consensus
exists that application to soil at a specific site is expected to sustainably
sequester carbon and concurrently improve soil functions (under current and
future management), while avoiding short- and long-term detrimental effects
to the wider environment as well as human and animal health.” As a material,
biochar is defined as: “charcoal for application to soil”.
The distinction between biochar as a concept and as a material is important.
For example, a particular biochar (material) may comply with all the conditions
in the concept of biochar when applied to field A, but not when applied to field
B. This report investigates the evidence for when, where and how actual
biochar application to soil complies with the concept, or not.
The terms ‘charcoal’ and ‘pyrogenic black carbon (BC)’ are also used in this
report when appropriate according to their definitions above and in the List of
Key Terms. Additionally, BC refers to C-rich residues from fire or heat
(including from coal, gas or petrol).
This report aims to review the state-of-the-art regarding the interactions
between biochar application to soil and its effects on soil properties,
processes and functioning. A number of recent publications have addressed
parts of this objective as well (Sohi et al., 2009; Lehmann and Joseph, 2009;
Collison et al., 2009). This report sets itself apart by i) addressing the issue
from an EU perspective, ii) inclusion of quantitative meta-analyses of selected
effects, and iii) a discussion of biochar for the threats to soil as identified by
the Thematic Strategy for Soil Protection (COM(2006) 231). In addition, this
h

report is independent, objective and critical.

Biochar is a stable carbon (C) compound created when biomass (feedstock)
is heated to temperatures between 300 and 1000ºC, under low (preferably
zero) oxygen concentrations. The objective of the biochar concept is to abate

31
the enhanced greenhouse effect by sequestering C in soils, while concurrently
improving soil quality. The proposed concept through which biochar
application to soils would lead to C sequestration is relatively straightforward.
Carbondioxide from the atmosphere is fixed in vegetation through
photosynthesis. Biochar is subsequently created through pyrolysis of the plant
material thereby potentially increasing its recalcitrance with respect to the
original plant material. The estimated residence time of biochar-carbon is in
the range of hundreds to thousands of years while the residence time of
carbon in plant material is in the range of decades. Consequently, this would
reduce the CO2 release back to the atmosphere if the carbon is indeed
persistently stored in the soil. The carbon storage potential of biochar is
widely hypothesised, although it is still largely unquantified, particularly when
also considering the effects on other greenhouse gasses (see Section 1.3),
and the secondary effects of large-scale biochar deployment. Concomitant
with carbon sequestration, biochar is intended to improve soil properties and
soil functioning relevant to agronomic and environmental performance.
Hypothesised mechanisms that have been suggested for potential
improvement are mainly improved water and nutrient retention (as well as
improved soil structure, drainage).
Considering the multi-dimensional and cross-cutting nature of biochar, an
imminent need is anticipated for a robust and balanced scientific review to
effectively inform policy development on the current state of knowledge with
reference to biochar application to soils.

How to read this report?

Chapter 1 introduces the concept of biochar and its origins, including a

comparison with European conditions/history.
Chapter 2 reviews the range of physical and chemical properties of biochars
that are most relevant to soils.
Chapter 3 focuses on the interactions between biochar application to soil
and soil properties, processes and functions.
Chapter 4 outlines how biochar application can be expected to influence
threats to soils.
Chapter 5 discusses some key issues regarding biochar that are beyond the
scope of this report.
Chapter 6 summarises the main findings of the previous chapters,
synthesises between these and identifies the key findings. Suggestions for
further reading are inserted where appropriate.

32
1.1 Biochar in the attention
The concept of biochar is increasingly in the attention in both political and
academic arenas, with several countries (e.g. UK, New Zealand, U.S.A.)
establishing ‘biochar research centres’; as well as in the popular media where
it is often portrayed as a miracle cure (or as a potential environmental
disaster). The attention of the media and public given to biochar can be
illustrated by contrasting a GoogleTM search for ‘biochar’ with a search for
‘biofuels’. A Google search for biochar yields 185,000 hits while biofuels yields
5,210,000 hits. Another illustration is given by comparing the search volumes
of ‘biochar’, ‘Terra Preta’ and ‘black earth’ over the last years, testifying the
recent increase in attention in and exposure of biochar (Figure 1.1, made with
Google TrendsTM).

14
Search Volume Index [-]

12 BIOCHAR
TERRA PRETA
10 BLACK EARTH

6
TM
Google Trends

0
Jul 17 2005

Jun 17 2007

Nov 4 2007
Jan 28 2007

Aug 10 2008
Jan 4 2004

May 23 2004

May 17 2009
Oct 10 2004

Feb 27 2005

Dec 4 2005

Apr 23 2006

Sep 10 2006

Mar 23 2008

Oct 4 2009
Dec 28 2008

Figure 1.1 Google TrendsTM result of “biochar”, “Terra Preta” and “black earth”. The scale is
based on the average worldwide traffic of “biochar” from January 2004 until June 2009 (search
performed on 04/12/2009)

The geographical interest in biochar can be explored further by using the

search volume index of biochar; the total number of searches normalised by
the overall search volume by country. Over the last 12 months the search
volume index for biochar was highest in Australia and New Zealand (Figure
1.2). The actual attention for biochar in Australia may even be higher, since in
Australia biochar is also referred to as ‘Agrichar’, one of its trade names.

33
Figure 1.2 Google TrendsTM geographical distribution of the search volume index of “biochar” of
the last 12 months from June 2008 to June 2009 (search performed on 16/09/2009). Data is
normalised against the overall search volume by country

An indication for the attention devoted to biochar by the scientific community

is provided by performing a search in the scientific literature search engines
Thompson’s ISI Web of Science and Google ScholarTM. A search in Google
ScholarTM yielded 724 hits for biochar and 48,600 hits for biofuels (searches
undertaken on 16/09/2009). If we consider ‘Terra Preta’ – a Hortic Anthrosol
found in Amazonia – in comparison to biochar, a search yielded 121,000 hits
on Google and 1,490 on Google Scholar. A search in the ISI Web of Science
for those articles indexed for either biochar or bio-char yielded a total of 81
articles (Figure 1.3). Three authors are independently involved in 22 articles
(~25%) of these 81 articles (Lehmann (9); Derimbas (8); Davaajav (8)). Out of
the 81 articles 27 articles include a reference to charcoal (Figure 1.3). This is
an indication of the relative small number of scientists currently involved in
biochar research, although the number of articles is rapidly increasing (Figure
1.3). Finally, the oldest paper appearing in either ISI Web of ScienceTM or
ScopusTM dealing with ‘biochar’, ‘Terra Preta’ or ‘black earth’ dates from 1998,
1984 and 1953, respectively.

34
 40
total biochar OR bio-char in ISI

35
biochar AND charcoal OR bio-char AND charcoal in ISI

30
Nr of articles in ISI (n)

25

20

15

10

0
1998 1999 2000 2001 2002 2003 2004 2005 2006 2007 2008 2009 2010

Figure 1.3 Scientific publications registred in Thompson’s ISI Web of Science indexed for either
biochar or bio-char including those articles that mention charcoal (search performed on
4/12/2009)

1.2 Historical perspective on soil improvement

Man-made soils (Anthrosols) enriched with charcoal are found as small
pockets (10s – 200 m in diameter) close to both current and historic human
settlements throughout Amazonia (see Figure 1.4) which are estimated to
cover a total area of 6,000 – 18,000 km2 (Sombroek and Carvalho de Souza,
2000). A rapidly expanding body of scientific literature has reached the
consensus that these soils were created by indigenous people, as far back as
10,000 yr BP (Woods et al., 2009), with varying depth (down to 1 m).

Figure 1.4 Distribution of Anthrosols in Amazonia (left; Glaser et al., 2001) and Europe (right;
Blume and Leinweber, 2004)

The first Anthrosols in Europe, which are mostly enriched with organic
material from peatlands and heathlands, have been dated to 3,000 yr BP on

35
the German island of Sylt (Blume and Leinweber, 2004). The largest expanse,
from a 3,500 km2 total European area of man-made soils (Plaggic
Anthrosols), was created during the Middle Ages in the nutrient poor, dry
sandy soils (Arenosols) of The Netherlands, northern Belgium and north-
western Germany (Figure 1.4) to similar depths as their Amazonian
counterparts (i.e. down to 1 m).
Such a vast single area of Anthrosols is rare, if not unique, and may be
explained by the relatively high population density (and subsequent food
demand) combined with environmental factors, i.e. the presence of extensive
peat deposits in close proximity to the nutrient poor free-draining soil. Much
more common are small scale Anthrosols, pockets of man-made soils close to
settlements, as an inevitable consequence or planned soil conditioning, by a
‘permanent’ human settlement that continuously produces organic waste.
Many Anthrosols do not appear on the EU soil distribution map because of
their small size in relation to the 1:1,000,000 scale of the Soil Geographical
Database of Eurasia, which is the basis of the map (Toth et al., 2008).
However, numerous small scale Anthrosols have been reported across the
European continent, e.g. Scotland (Meharg et al., 2006; Davidson et al.,
2006), Ireland, Italy, Spain and northwest Russia (Giani et al., 2004). Based
on their formation, it can be assumed that Anthrosols exist in other parts of
Europe as well, but data are lacking.

Figure 1.5 Comparing tropical with temperate Anthrosols. The left half shows a profile of a
fertile Terra Preta (Anthrosol with charcoal) created by adding charcoal to the naturally-
occurring nutrient poor Oxisol (far left; photo courtesy of Bruno Glaser). The right half (far
right) is a profile picture of a fertile European Plaggen Soil (Plaggic Anthrosol; photo courtesy of
Erica Micheli) created by adding peat and manure to the naturally-occurring nutrient poor
sandy soils (Arenosols) of The Netherlands

Although both European and Amazonian Anthrosols were enriched to

increase their agricultural performance, there is an important distinction
between the Plaggic Anthrosols of Europe and the Hortic Anthrosols of
Amazonia (Figure 1.5). Plaggic is from the Dutch ‘Plag’ meaning a cut out
section of the organic topsoil layer, including vegetation (grass or heather)
while Hortic Anthrosol translates freely into ‘kitchen soil’. These names are
reflected in their composition, i.e. Plaggic Anthrosols were made by adding
organic topsoil material and peat (early Middle Ages) and mixed with manure
(late Middle Ages) while Hortic Anthrosols were created by a wide variety of
organic and mineral materials, ranging from animal bones to charcoal and
pottery fragments. What sets the Terra Preta apart from other Hortic

36
Anthrosols is the high proportion of charcoal. It is assumed that the charcoal
was made deliberately for application to soil, i.e. not just charred remains from
clearing and burning the forest.

1.3 Different solutions to similar problems

The challenges faced by the people of two very different environments
(tropical rain forest vs. temperate climate on largely open or partially
deforested land) appear similar in the sense of needing to grow crops on soils
that naturally had low nutrient and water retention. One can only speculate as
to what exactly the reasons were for the people living at the time to either add
or not add charcoal to their soils. In addition to the available supply of organic
materials, possible explanations may be related to the relative value of the
different organic materials and contrasting residence times of SOM. In a
simplified scenario, the colder climate in Europe means that microbial
decomposition occurs much more slowly than in the tropics, leading to much
longer residence times of organic matter. The recalcitrance of the peat and
plaggen that were added to the soil meant that the benefits from increased
water and (to a lesser degree) nutrient retention lasted long enough to make it
worth the investment. In tropical soils, however, the recalcitrance of the
organic matter that was added to the soil needed to be greater to get a return
that was worth the investment. Charring organic matter may have been a
conscious policy to achieve this. Of course, wood and charcoal were being
produced in Europe at the time as well. However, other uses of these
materials were likely to be more valuable, e.g. the burning of wood in fire
places to heat living accommodations and the use of charcoal to achieve high
enough temperatures for extracting metals from ores.
Because of the relatively small areal extent of Anthrosols, many of their
locations may not be known or recognised presently. It is possible that small
pockets of Anthrosols exist in Europe, created at different times in history,
where greater amounts of charcoal are present than in the Plaggic Anthrosols.
Potentially, identification and study of these sites (including chronosequences)
could provide valuable information regarding the interactions between
charcoal and environmental factors prevalent in Europe.

1.4 Biochar and pyrogenic black carbon

A potential analogue for biochar may be found in the charcoal produced by
wildfires (or pyrogenic black carbon – BC – as it is often referred to) found
naturally in soils across the world, and in some places even makes up a larger
proportion of total organic C in the soil than in some Terra Preta soils. Preston
and Schmidt (2006) showed an overview of studies on non-forested sites in
different parts of the world with BC making up between 1 and 80% of total
SOC. For example, BC was found to constitute 10-35% of the total SOC
content for five soils from long-term agricultural research sites across the
U.S.A. (Skjemstad et al., 2002). Schmidt et al. (1999) studied pyrogenic BC
contents of chernozemic soils (Cambisol, Luvisol, Phaeozem, Chernozem and
Greyzem) in Germany and found BC to make up 2-45% of total SOC (mean of
14%).

37
Figure 1.6 Terms and properties of pyrogenic BC (adopted from Preston and Schmidt, 2006)

However, it is important to bear in mind that, while the range of BC materials

produced by wildfire overlaps with the range of biochar materials (i.e. the
continuum from charred biomass to soot and graphite; Figure 1.6), the
composition and properties of biochar can be very different to pyrogenic BC
(see Chapter 2). The two main responsible factors are feedstock and pyrolysis
conditions. In a wildfire, the feedstock is the aboveground biomass (and
sometimes peat and roots) while for biochar any organic feedstock can
theoretically be used from wood and straw to chicken manure (Chapter 2). In
a pyrolysis oven, the pyrolysis conditions can be selected and controlled,
including maximum temperature and duration but also the rate of temperature
increase, and inclusion of steam, or e.g. KOH, activation and oxygen
conditions.

1.5 Carbon sequestration potential

Globally, soil is estimated to hold more organic carbon (1,100 Gt; 1
Gt=1,000,000,000 tonnes) than the atmosphere (750 Gt) and the terrestrial
biosphere (560 Gt) (Post et al., 1990; Sundquist, 1993). In the Kyoto Protocol
on Climate Change of 1997, which was adopted in the United Nations
Framework Convention on Climate Change, Article 3.4 allows organic carbon
stored in arable soils to be included in calculations of net carbon emissions. It
speaks of the possibility of subtracting the amounts of CO2 removed from the
atmosphere into agricultural sinks, from the assigned target reductions for
individual countries. SOC sequestration in arable agriculture has been
researched (Schlesinger, 1999; Smith et al., 2000a, b; Freibauer et al., 2002;
West & Post, 2002; Sleutel et al., 2003; Janzen, 2004; King et al., 2004; Lal,
2004) against the background of organic carbon (OC) credit trading schemes
(Brown et al., 2001; Johnson & Heinen, 2004). However, fundamental
knowledge on attainable SOC contents (relative to variation in environmental
factors) is still in its infancy, and it is mostly approached by modelling (Falloon
et al., 1998; Pendall et al., 2004).
The principle of using biochar for carbon (C) sequestration is related to the
role of soils in the C-cycle (Figure 1.7). As Figure 1.7 shows, the global flux of
CO2 from soils to the atmosphere is in the region of 60 Gt of C per year. This
CO2 is mainly the result of microbial respiration within the soil system as the
microbes decompose soil organic matter (SOM). Components of biochar are
proposed to be considerably more recalcitrant than SOM and as such are only
decomposed very slowly, over a time frame which can be measured in

38
hundreds or thousands of years. This means that biochar allows carbon input
into soil to be increased greatly compared to the carbon output through soil
microbial respiration, and it is this that is the basis behind biochar’s possible
carbon negativity and hence its potential for climate change mitigation.

Figure 1.7 Diagram of the carbon cycle. The black numbers indicate how much carbon is stored
in various reservoirs, in billions of tons (GtC = Gigatons of Carbon and figures are circa 2004).
The purple numbers indicate how much carbon moves between reservoirs each year, i.e. the
fluxes. The sediments, as defined in this diagram, do not include the ~70 million GtC of
carbonate rock and kerogen (NASA, 2008)

Although Figure 1.7 is clearly a simplification of the C-cycle as it occurs in

nature, the numbers are well established (NASA, 2008) and relatively
uncontroversial. A calculation of the fluxes, while being more a ‘back of the
envelope’ calculation, than precise mathematics, is highly demonstrative of
the anthropogenic influence on atmospheric CO2 levels. When all of the sinks
are added together (that is the fluxes of CO2 leaving the atmosphere) the total
amount of C going into sinks is found to be in the region of 213.35 Gt per
year. Conversely, when all of the C fluxes emitted into the atmosphere from
non-anthropogenic (natural) sources are added, they total 211.6 Gt per year.
This equates to a net loss of carbon from the atmosphere of 1.75 Gt C.
It is for this reason that the relatively small flux of CO2 from anthropogenic
sources (5.5 Gt C per year) is of such consequence as it turns the overall C
flux from the atmosphere from a loss of 1.75 Gt per year, to a net gain of 3.75
Gt C per year. This is in relatively close agreement with the predicted rate of
CO2 increase of about 3 Gt of C per year (IPCC, 2001). It is mitigation of this
net gain of CO2 to the atmosphere that biochar’s addition to soil is posited for.
Lehmann et al. (2006) estimate a potential global C-sequestration of 0.16 Gt
yr-1 using current forestry and agricultural wastes, such as forest residues, mill
residues, field crop residues, and urban wastes for biochar production. Using

39
projections of renewable fuels by 2100, the same authors estimate
sequestration to reach a potential range of 5.5-9.5 Gt yr-1, thereby exceeding
current fossil fuel emissions. However, the use of biochar for climate change
mitigation is beyond the scope of this report that focuses on the effects of
biochar addition to soils with regard to physical, chemical and biological
effects, as well as related effects on soil and ecosystem functioning.

1.5.1 Biochar loading capacity

Terra Preta soils have been shown to contain about 50 t C ha-1 in the form of
BC, down to a depth of approximately 1 meter (approximately double the
amount relative to pre-existing soil), and these soils are highly fertile when
compared to the surrounding soils. This has lead to the idea of biochar being
applied to soil to sequester carbon and maintain or improve the soil
production function (e.g. crop yields), as well as the regulation function and
habitat function of soils. Controlled experiments have been undertaken to look
at the effects of different application rates of biochar to soils.
At present, however, it is not clear whether there is a maximum amount of C,
in the form of biochar, which can be safely added to soils without
compromising other soil functions or the wider environment; that is, what is
the ‘biochar loading capacity’ (BLC) of a given soil? It will be important to
determine if the BLC varies between soil types and whether it is influenced by
the crop type grown on the soil. In order to maximise the amount of biochar
which can be stored in soils without impacting negatively on other soil
functions, the biochar loading capacity of different soils exposed to different
environmental and climatic conditions specific to the site will have to be
quantified for different types of biochar.
The organic matter fractions of some soils in Europe have been reported to
consist of approximately 14% (up to 45%) BC or charcoal (see Section 1.4),
which are both analogues of biochar as previously discussed. Lehmann and
Rondon (2005) reported that at loadings up to 140 t C ha-1 (in a weathered
tropical soil) positive yield effects still occurred. However, it should be noted
that some experiments report that some crops experience a loss of the
positive effects of biochar addition to soil at a much lower application rate. For
example, Rondon et al. (2007) reported that the beans (Phaseolus vulgaris L.)
showed positive yield effects on biochar application rates up to 50 t C ha-1 that
disappeared at an application rate of 60 t C ha-1 with a negative effect on yield
being reported at application rates of 150 t C ha-1. This shows that the BLC is
likely to be crop dependent as well as probably both soil and climate
dependent. Combined with the irreversibility of biochar application to soil, this
highlights the complex nature of calculating a soil’s BLC as future croppings
should be taken into account to ensure that future crop productivity is not
compromised if the crop type for a given field is changed. Apart from effects
on plant productivity, it can be imagined that other effects, on for example soil
biology or transport of fine particles to ground and surface water, should be
taken into account when ‘calculating’ or deriving the BLC for a specific site.
Also, the BLC concept would need to be developed for both total (final)
amount and the rate of application, i.e. the increase in the total amount over
time. The rate of application would need to consist of a long term rate (i.e. t
ha-1 yr-1 over 10 or 100 years) as well as a ‘per application’ rate, both

40
determined by evidence of direct and indirect effects on soil and the wider
environment.
Another consideration regarding the biochar loading capacity of a soil is the
risk of smouldering combustion. Organic soils that dry out sufficiently are
capable of supporting below ground smouldering combustion that can
continue for long time periods (years in some cases). It is feasible that soils
which experience very high to extreme loading rates of biochar and are
subject to sufficiently dry conditions could support smouldering fires. Ignition
of such fires could occur both naturally, e.g. by lightening strike, or
anthropogenically. What the biochar content threshold would be, how the
threshold would change according to environmental conditions, and how
much a risk this would be in non-arid soils remains unclear, but is certainly
worthy of thought and future investigation.

1.5.2 Other greenhouse gasses

Carbon dioxide is not the only gas emitted from soil with the potential to
influence the climate. Methane (CH4) production also occurs as a part of the
carbon cycle. It is produced by the soil microbiota under anaerobic conditions
through a process known as methanogenesis and is approximately 21 times
more potent as a greenhouse gas than CO2 over a time horizon of 100 years.
Nitrous oxide (N2O) is produced as a part of the nitrogen (N) cycle through
process known as nitrification and denitrification which are carried out by the
soil microbiota. Nitrous oxide is 310 times more potent as a greenhouse gas
than CO2over a time horizon of 100 years (U.S. Environmental Protection
Agency, 2002).
Whilst these gases are more potent greenhouse gases than CO2, only
approximately 8% of emitted greenhouse gases are CH4 and only 5% are
N2O, with CO2 making up approximately 83% of the total greenhouse gases
emitted. Eighty percent of N20 and 50% of CH4 emitted are produced by soil
processes in managed ecosystems (US Environmental Protection Agency,
2002). It should be noted that these figures detail total proportions of each
greenhouse gas and are not weighted to account for climatic forcing.
In one study, biochar addition to soils has been shown to reduce the emission
of both CH4 and N2O. Rondon et al. (2005) reported that a near complete
suppression of methane upon biochar addition at an application rate of 2% w
w-1 to soil. It was hypothesised that the mechanism leading to reduced
emission of CH4 is increased soil aeration leading to a reduction in frequency
and extent of anaerobic conditions under which methanogenesis occurs.
Pandolfo et al. (1994) investigated CH4 adsorption capacity of several
activated carbons (from coconut feedstock) in a series of laboratory
experiments. Their results showed increased CH4 ‘adsoprtion’ with increase
surface area of the activated carbon, particularly for micropores (<2µm).
These charcoal materials were activated using steam or KOH, however, and it
remains to be tested how different biochar materials added to soils in the field
will interact with methane dynamics. The influence of biochar on SOM
dynamics are discussed later in this report (Section 3.2.5).
A reduction in N2O emissions of 50% in soybean plantations and 80% in
grass stands was also reported (Rondon et al. 2005). The authors

41
hypothesised that the mechanism leading to this reduction in N2O emissions
was due to slower N cycling, possibly as a result of an increase in the C:N
ratio. It is also possible that the N that exists within the biochar is not
bioavailable when introduced to the soil as it is bound up in heterocyclic form
(Camps, 2009; Personal communication). Yanai et al. (2007) measured N2O
emissions from soils after rewetting in the laboratory and found variable
results, i.e. an 89% suppression of N2O emissions at 73-78% water-filled pore
space contrasting to a 51% increase at 83% water-filled pore space. These
results indicate that the effect of biochar additions to soils on the N cycle
depend greatly on the associated changes in soil hydrology and that
thresholds of water content effects on N20 production may be very important
and would have to be studied for a variety of soil-biochar-climate conditions.
Furthermore, if biochar addition to soil does slow the N-cycle, this could have
possible consequences on soil fertility in the long term. This is because nitrate
production in the soil may be slowed beyond the point of plant uptake,
meaning that nitrogen availability, often the limiting factor for plant growth in
soils, may be reduced leading to concurrent reduction in crop productivity.
Yanai et al. (2007) reported that this effect did change over time, but their
experiment only ran for 5 days and so extrapolation of the results to the time
scales at which biochar is likely to persist in soil is not possible. Further
research is therefore needed to better elucidate the effects and allow
extrapolation to the necessary time scales.

1.6 Pyrolysis
Pyrolysis is the chemical decomposition of an organic substance by heating in
the absence of oxygen. The word is derived from Greek word ‘pyro’ meaning
fire and “lysis” meaning decomposition or breaking down into constituent
parts. In practice it is not possible to create a completely oxygen free
environment and as such a small amount of oxidation will always occur.
However, the degree of oxidation of the organic matter is relatively small
when compared to combustion where almost complete oxidation of organic
matter occurs, and as such a substantially larger proportion of the carbon in
the feedstock remains and is not given off as CO2. However, with pyrolysis
much of the C from the feedstock is still not recovered in charcoal form, but
converted to either gas or oil.
Pyrolysis occurs spontaneously at high temperatures (generally above
approximately 300°C for wood, with the specific temperature varying with
material). It occurs in nature when vegetation is exposed to wildfires or comes
into contact with lava from volcanic eruptions. At its most extreme, pyrolysis
leaves only carbon as the residue and is called carbonization. The high
temperatures used in pyrolysis can induce polymerisation of the molecules
within the feedstocks, whereby larger molecules are also produced (including
both aromatic and aliphatic compounds), as well as the thermal
decomposition of some components of the feedstocks into smaller molecules.
This is discussed in more detail in Section 3.2.5.1.
The process of pyrolysis transforms organic materials into three different
components, being gas, liquid or solid in different proportions depending upon
both the feedstock and the pyrolysis conditions used. Gases which are
produced are flammable, including methane and other hydrocarbons which

42
can be cooled whereby they condense and form an oil/tar residue which
generally contains small amounts of water. The gasses (either condenses or
in gaseous form) and liquids can be upgraded and used as a fuel for
combustion.
The remaining solid component after pyrolysis is charcoal, referred to as
biochar when it is produced with the intention of adding it to soil to improve it
(see List of Key terms). The physical and chemical properties of biochar are
discussed in more detail in Chapter 2.

The process of pyrolysis has been adopted by the chemical industry for the
production of a range of compounds including charcoal, activated carbon,
methanol and syngas, to turn coal into coke as well as producing other
chemicals from wood. It is also used for the breaking down, or ‘cracking’ of
medium-weight hydrocarbons from oil to produce lighter hydrocarbons such
as petrol.
A range of compounds in the natural environment are produced by both
anthropogenic and non-anthropogenic pyrolysis. These include compounds
released from the incomplete burning of petrol and diesel in internal
combustion engines, through to particles produced from wood burned in forest
fires, for example. These substances are generally referred to as black carbon
(see List of Key terms) in the scientific literature and exist in various forms
ranging form small particulate matter found in the atmosphere, through to a
range of sizes found in soils and sediments where it makes up a significant
part of the organic matter (Schmidt et al., 1999; Skjemstad et al., 2002;
Preston et al., 2006; Hussain et al. 2008).

1.6.1 The History of Pyrolysis

While it is possible that pyrolysis was first used to make charcoal over 7,000
years ago for the smelting of copper, or even 30,000 years ago for the
charcoal drawings of the Chauvet cave (Antal, 2003), the first definitive
evidence of pyrolysis for charcoal production comes from over 5,500 years
ago in Southern Europe and the Middle East. By 4,000 years ago, the start of
the Bronze Age, pyrolysis use for the production of charcoal must have been
widespread. This is because only burning charcoal allowed the necessary
temperatures to be reached to smelt tin with copper and so produce bronze
(Earl, 1995).
A range of compounds can be found in the natural environment that is
produced by both anthropogenic and non-anthropogenic pyrolysis. These
include compounds released from the incomplete burning of petrol and diesel
in internal combustion engines, through to being produced from wood in forest
fires for example.

1.6.2 Methods of Pyrolysis

Although the basic process of pyrolysis, that of heating a C-containing
feedstock in an limited oxygen environment, is always the same, different
methodologies exist, each with different outputs.
Apart from the feedstocks used, which are discussed further is Section 1.7,
the main variables that are often manipulated are pyrolysis temperature, and

43
the residence time of the feedstock in the pyrolysis unit. Temperature itself
can have a large effect on the relative proportions of end product from a
feedstock (Fig. 1.9).

80
Biochar
70 Biooil
Gas
60
Water
50
Yield (%wt)

40

30

20

10

0
400 450 500 550 600 650

Temperature (°C)

Figure 1.8 A graph showing the relative proportions of end products after fast pyrolysis of aspen
poplar at a range of temperatures (adapted from IEA, 2007)

Residence times of both the solid constituents and the hot vapor produced
under pyrolysis conditions can also have a large effect on the relative
proportions of each end product of pyrolysis (Table 1.1). In the nomenclature,
four different types of pyrolysis are generally referred to, with the difference
between each being dependent on temperature and residence time of solid or
vapour in the pyrolysis unit, or a combination of both. The four different types
of pyrolysis are fast, intermediate and slow pyrolysis (with slow pyrolysis often
referred to as “carbonisation” due to the relatively high proportion of
carbonaceous material it produces: biochar) along with gasification (due to the
high proportion of syngas produced).

Table 1.1 shows that different pyrolysis conditions lead to different proportions
of each end product (liquid, char or gas). This means that specific pyrolysis
conditions can be tailored to each desired outcome. For example, the IEA
report (2007) stated that fast pyrolysis was of particular interest as liquids can
be stored and transported more easily and at lower cost than solid or gaseous
biomass forms. However, with regard to the use of biochar as a soil
amendment and for climate change mitigation it is clear that slow pyrolysis,
would be preferable, as this maximises the yield of char, the most stable of
the pyrolysis end products.

44
Table 1.1 The mean post-pyrolysis feedstock residues resulting from different temperatures and
residence times (adapted from IEA, 2007)

Mode Conditions Liquid Biochar Syngas

Moderate temperature, ~500°C, short hot

Fast pyrolysis 75% 12% 13%
vapour residence time of ~ 1 s
Intermediate Moderate temperature ~500°C, moderate
50% 20% 30%
Pyrolysis hot vapour residence time of 10 – 20 s
Slow Pyrolysis Low temperature ~400°C,
30% 35% 35%
(Carbonisation) very long solids residence time
High temperature ~800°C,
Gasification 5% 10% 85%
long vapour residence time

Owing to the fact that end products such as flammable gas can be recycled
into the pyrolysis unit and so provide energy for subsequent pyrolysis cycles,
costs, both in terms of fuel costs, and of carbon emission costs, can be
minimised. Furthermore, the pyrolysis reaction itself becomes exothermic
after a threshold is passed, thereby reducing the required energy input to
maintain the reaction. However, it is important to note that other external costs
are associated with pyrolysis, most of which will be discussed in Section 2.4.
For example, fast pyrolysis requires that the feedstock is dried to less than
10% water (w w-1). This is done so that the bio-oil is not contaminated with
water. The feedstock then needs to be ground to a particle size of ca. 2 mm to
ensure that there is sufficient surface area to ensure rapid reaction under
pyrolysis conditions (IEA, 2007). The grinding of the feedstock, and in some
cases also the drying require energy input and will increase costs, as well as
of the carbon footprint of biochar production if the required energy is not
produced by carbon neutral sources.
As well as different pyrolysis conditions, the scale at which pyrolysis is
undertaken can also vary greatly. The two different scales discussed
throughout this report are that of ‘Closed’ vs ‘Open’ scenarios. Closed refers
to the scenario in which relatively small, possibly even mobile, pyrolysis units
are used on each farm site, with crop residues and other bio-wastes being
pyrolysed on site and added back to the same farm’s soils. Open refers to
biowastes being accumulated and pyrolysed off-site at industrial scale
pyrolysis plants, before the biochar is redistributed back to farms for
application to soil. The scales at which these scenarios function are very
different, and each brings its own advantages and disadvantages.

1.7 Feedstocks
Feedstock is the term conventionally used for the type of biomass that is
pyrolysed and turned into biochar. In principle, any organic feedstock can be
pyrolysed, although the yield of solid residue (char) respective to liquid and
gas yield varies greatly (see Section 1.6.2) along with physico-chemical
properties of the resulting biochar (see Chapter 2).
Feedstock is, along with pyrolysis conditions, the most important factor
controlling the properties of the resulting biochar. Firstly, the chemical and

45
structural composition of the biomass feedstock relates to the chemical and
structural composition of the resulting biochar and, therefore, is reflected in its
behaviour, function and fate in soils. Secondly, the extent of the physical and
chemical alterations undergone by the biomass during pyrolysis (e.g. attrition,
cracking, microstructural rearrangements) are dependent on the processing
conditions (mainly temperature and residence times). Table 1.2 provides a
summary of some of the key components in representative biochar
feedstocks.

Table 1.2 Summary of key components (by weight) in biochar feedstocks (adapted from Brown
et al., 2009)

Ash Lignin Cellulose

-1
(w w )
Wheat straw 11.2 14 38
Maize residue 2.8-6.8 15 39
Switchgrass 6 18 32
Wood (poplar, 0.27 - 1 26 - 30 38 - 45
willow, oak)

Cellulose and ligning undergo thermal degradation at temperatures ranging

between 240-350ºC and 280-500ºC, respectively (Sjöström, 1993; Demirbas,
2004). The relative proportion of each component will, therefore, determine
the extent to which the biomass structure is retained during pyrolysis, at any
given temperature. For example, pyrolysis of wood-based feedstocks
generates coarser and more resistant biochars with carbon contents of up to
80%, as the rigid ligninolytic nature of the source material is retained in the
biochar residue (Winsley, 2007). Biomass with high lignin contents (e.g. olive
husks) have shown to produce some of the highest biochar yields, given the
stability of lignin to thermal degradation, as demonstrated by Demirbas
(2004). Therefore, for comparable temperatures and residence times, lignin
loss is tipically less than half of cellulose loss (Demirbas, 2004).
Whereas woody feedstock generally contains low proportions (< 1% by
weight) of ash, biomass with high mineral contents such as grass, grain husks
and straw residues generally produce ash-rich biochar (Demirbas 2004).
These latter feedstocks may contain ash up to 24% or even 41% by weight,
such as rice husk (Amonette and Joseph, 2009) and rice hulls (Antal and
Grønly, 2003), respectively. The mineral content of the feedstock is largely
retained in the resulting biochar, where it concentrates due to the gradual loss
of C, hydrogen (H) and oxygen (O) during processing (Demirbas 2004). The
mineral ash content of the feedstock can vary widely and evidence seems to
suggest a relationship between that and biochar yield (Amonette and Joseph,
2009). Table 1.3 provides an example of the elemental composition of
representative feedstocks.

46
Table 1.3 Examples of the proportions of nutrients (g kg-1) in feedstocks (adapted from Chan
and Xu, 2009)

Ca Mg K P
-1
(g kg )
Wheat straw 7.70 4.30 2.90 0.21
Maize cob 0.18 1.70 9.40 0.45
Maize stalk 4.70 5.90 0.03 2.10
Olive kernel 97.0 20.0 - -
Forest residue 130 19.0 - -

In the plant, Ca occurs mainly within cell walls, where it is bound to organic
acids, while Mg and P are bound to complex organic compounds within the
cell (Marschner, 1995). Potassium is the most abundant cation in higher
plants and is involved in plant nutrition, growth and osmoregulation
(Schachtman and Schroeder, 1994). Nitrogen, Mn and Fe also occur
associated to a number of organic and inorganic forms. During thermal
degradation of the biomass, potassium (K), chlorine (Cl) and N vaporize at
relatively low temperatures, while calcium (Ca), magnesium (Mg), phosphorus
(P) and sulphur (S), due to increased stability, vaporise at temperatures that
are considerably higher (Amonette and Joseph, 2009). Other relevant
minerals can occur in the biomass, such as silicon (Si), which occurs in the
cell walls, mostly in the form of silica (SiO2).
Many different materials have been proposed as biomass feedstocks for
biochar, including wood, grain husks, nut shells, manure and crop residues,
while those with the highest carbon contents (e.g. wood, nut shells),
abundancy and lower associated costs are currently used for the production
of activated carbon (e.g. Lua et al., 2004; Martinez et al., 2006; Gonzaléz et
al., 2009;). Other feedstocks are potentially available for biochar production,
among which biowaste (e.g. sewage sludge, municipal waste, chicken litter)
and compost. Nevertheless, a risk is associated to the use of such source
materials, mostly linked to the occurrence of hazardous components (e.g.
organic pollutants, heavy metals). Crystalline silica has also been found to
occur in some biochars. Rice husk and rice straw contain unusually high
levels of silica (220 and 170 g kg-1) compared to that in other major crops.
High concentrations of calcium carbonate (CaCO3) can be found in pulp and
paper sludge (van Zwieten et al., 2007) and are retained in the ash fraction of
some biochars.
Regarding the characteristics of some plant feedstocks, Collison et al. (2009)
go further, suggesting that even within a biomass feedstock type, different
composition may arise from distinct growing environmental conditions (e.g.
soil type, temperature and moisture content) and those relating to the time of
harvest. In corroboration, Wingate et al. (2009) have shown that the adsorbing
properties of a charcoal for copper ions can be improved 3-fold by carefully
selecting the growth conditions of the plant biomass (in this case, stinging
nettles). Even within the same plant material, compositional heterogeneity has

47
also been found to occur among different parts of the same plant (e.g. maize
cob and maize stalk, Table 1.3).
Lignocellulosic biomass is an obvious feedstock choice because it is one of
the most abundant naturally occurring available materials (Amonette and
Joseph, 2009). The spatio-temporal occurrence of biomass feedstock will
influence the availability of specific biochars and its economic value (e.g.
distance from source to field). For example, in an area with predominantly root
crops on calcareous sandy arable soils and a dry climate, biochars that
provide more water retention and are mechanically strong (e.g. woody
feedstocks) are likely to be substantially more valuable than in an area of
predominantly combinable crops on acidic sandy soils and a ‘year round’ wet
climate. In the latter case, biochars with a greater CEC, liming capacity and
possibly a lower mechanical strength (e.g. crop residue feedstock) may be
more in demand.
In Terra Pretas potential feedstocks were limited to wood from the trees and
organic matter from other vegetation. Nowadays any biomass material,
including waste, is considered as a feedstock for biochar production.
Considering that historical sites contain either biochar (Terra Preta) or BC
(from wildfires), chronosequence studies can only give us information about
the long term consequences and dynamics of those limited natural
feedstocks. This implies an important methodological challenge for the study
of the long term dynamics of soils with biochar produced from feedstocks
other than natural vegetation. Even for trees and plants, careful consideration
needs to be given to specific species that bioaccumulate certain metals, or, in
the case of crop residues, that may contain relevant concentrations of
herbicides, pesticides, fungicides, and in the case of animal manures that may
contain antibiotics or their secondary metabolites. See Section 5.1.5 for a
more detailed discussion on the (potential) occurrence of contaminants within
biochar.
In addition, chronosequence studies using historic sites are often poor
predictors of structural disintegration and concomitant chemical reactivity and
mobility of biochars, because they are either not in arable land use, or have
not been subject to the intense physical disturbance of modern arable tillage
and cultivation (e.g. the power harrow).
A detailed description of all biochar feedstocks is beyond the scope of this
report and feedstocks have been reviewed in other works (Collison et al.,
2009; Lehmann and Joseph, 2009). The key point is that the suitability of
each biomass type as a potential source for biochar, is dependent on a
number of chemical, physical, environmental, as well as economic and
logistical factors (Collison et al., 2009), as discussed, where appropriate,
throughout this report. It is important to stress, however, that for any material
to be considered as a feedstock for biochar production, and therefore also for
application to soil, a rigorous procedure needs to be developed in order to
assess the biochar characteristics and long term dynamics in the range of
soil, other environmental conditions, and land use and management factors
that are considered for its application.

48
1.8 Application Strategies
Biochar application strategies have been studied very little, although the way
biochar is applied to soils can have a substantial impact on soil processes and
functioning, including aspects of the behaviour and fate of biochar particles in
soil and the wider environment (Chapter 3) as well as on ‘threats to soil’
(Chapter 4), occupational health and safety (5.2), and economic
considerations (Section 5.4). Broadly speaking there are three main
approaches: i) topsoil incorporation, ii) depth application, and iii) top-dressing.

For topsoil incorporation biochar can be applied on its own or combined with
composts or manures. The degree of mixing will depend on the cultivation
techniques used. In conventional tillage systems the biochar (and
compost/manure/slurry) will generally be mixed more or less homogeneously
throughout the topsoil (in most arable soils from 0-15/30 cm depth). Water
and wind erosion will remove biochar along with other soil material, i.e. that
would erode without biochar additions as well, and possibly more biochar will
be eroded from the surface because of its low density. Potentially, the
application of biochar combined with compost or manure would reduce this
risk, but studies evidencing this are lacking. In conservation tillage systems
the incorporation depth will be reduced (leading to greater biochar
concentrations at equal application rates) and possibly a concentration
gradient decreasing with depth. In no-till systems any incorporation would be
through natural processes (see top-dressing below). Deep mouldboard
ploughing effectively results in (temporary) ‘depth application’ (see below),
with more topsoil homogenisation occurring during subsequent ploughing.

Depth application of biochar has been described mostly as ‘deep-banded’

application (e.g. Blackwell et al., 2007). The placement of the biochar directly
into the rhizosphere is thought to be more beneficial for crop growth and less
susceptible to erosion. The application can be either by pneumatic systems,
which can operate at high rates, or by applying the biochar in furrows or
trenches and subsequently levelling the soil surface. Deep mouldboard
ploughing essentially results in temporary ‘depth application’, although
horizontally continuous (unlike the ‘deep-banded’ application). Subsequent
mouldboard ploughing and cultivation will then further homogenise the biochar
distribution through the topsoil.

Top-dressing of biochar is the spreading of biochar (dust fraction mostly) to

the soil surface and relying on natural processes for the incorporation of the
biochar into the topsoil. This form of application is being considered mainly for
those situations where mechanical incorporation is not possible, e.g. no-till
systems, forests, and pastures. An obvious drawback is the risk of erosion by
water and wind, as well as human health (inhalation) and impacts on other
ecosystem components (e.g. surface water, leaf surfaces, etc.). It is also
largely unknown what the rates of incorporation would be for different soil-
climate-land use combinations.

The dust fraction of biochar is an issue for all application strategies during the
storaging, handling, and applying phases of the biochar (see Sections 2.2.1
and 5.2 for more detailed information about the properties and implications of

49
biochar’s dust fraction).This aspects needs to be investigated thoroughly
before implementation. Like any trafficking on soil, there is a risk of (sub)soil
compaction during biochar application. This may be particularly the case for
the relatively heavy machinery involved in ‘depth application’.

Both topsoil incorporation and top-dressing can be applied with a range of

frequencies, i.e. a ‘one-off’ application’, every few years, or every year. For
specific effects on soil, e.g. nutrient availability (from a feedstock like poultry
manure) or liming effect, a more frequent application may be more beneficial
to the soil and/or less detrimental to the environment (nitrate leaching).

1.9 Summary
As a concept biochar is defined as ‘charcoal (biomass that has been
pyrolysed in a zero or low oxygen environment) for which, owing to its
inherent properties, scientific consensus exists that application to soil at a
specific site is expected to sustainably sequester carbon and concurrently
improve soil functions (under current and future management), while avoiding
short- and long-term detrimental effects to the wider environment as well as
human and animal health'. Inspiration is derived from the anthropogenically
created Terra Preta soils (Hortic Anthrosols) in Amazonia where charred
organic material plus other (organic and mineral) materials appear to have
been added purposefully to soil to increase its agronomic quality. Ancient
Anthrosols have been found in Europe as well, where organic matter (peat,
manure, ‘plaggen’) was added to soil, but where charcoal additions appear to
have been limited or non-existent. Furthermore, charcoal from wildfires
(pyrogenic black carbon - BC) has been found in many soils around the world,
including European soils where pyrogenic BC can make up a large proportion
of total soil organic carbon.
Biochar can be produced from a wide range of organic feedstocks under
different pyrolysis conditions and at a range of scales. Many different
materials have been proposed as biomass feedstocks for biochar. The
suitability of each biomass type for such an application is dependent on a
number of chemical, physical, environmental, as well as economic and
logistical factors. The original feedstock used, combined with the pyrolysis
conditions will determine the properties, both physical and chemical, of the
biochar product. It is these differences in physicochemical properties that
govern the specific interactions which will occur with the endemic soil biota
upon addition of biochar to soil, and hence how soil dependent ecosystem
functions and services are affected. The application strategy used to apply
biochar to soils is an important factor to consider when evaluating the effects
of biochar on soil properties and processes. Furthermore, the biochar loading
capacity of soils has not been fully quantified, or even developed
conceptually.

50
2. PHYSICOCHEMICAL PROPERTIES OF BIOCHAR
This chapter provides an overview of the physical and chemical properties of
biochar, as determined mainly by feedstock and the pyrolysis operational
conditions. The combined heterogeneity of the feedstock and the wide range
of chemical reactions which occur during processing, give rise to a biochar
product with a unique set of structural and chemical characteristics (Antal and
Gronli, 2003; Demirbas, 2004). A primary focus was given to those
characteristics that are more likely to impact on soil properties and processes
when biochar is incorporated into soil. The implications of such characteristics
in the context of the biochar-soil mixture are discussed in Chapter 3. More
detailed information on a wider range of biochar properties can be found in
the relevant scientific literature (e.g. Lehmann and Joseph, 2009; and others).

2.1 Structural and Chemical Composition

2.1.1 Structural composition

Thermal degradation of cellulose between 250 and 350ºC results in
considerable mass loss in the form of volatiles, leaving behind a rigid
amorphous C matrix. As the pyrolysis temperature increases, so thus the
proportion of aromatic carbon in the biochar, due to the relative increase in
the loss of volatile matter (initially water, followed by hydrocarbons, tarry
vapours, H2, CO and CO2), and the conversion of alkyl and O-alkyl C to aryl C
(Baldock and Smernik, 2002; Demirbas 2004). Around 330ºC, polyaromatic
graphene sheets begin to grow laterally, at the expense of the amorphous C
phase, and eventually coalesce. Above 600ºC, carbonization becomes the
dominant process. Carbonization is marked by the removal of most remaining
non-C atoms and consequent relative increase of the C content, which can be
up to 90% (by weight) in biochars from woody feedstocks (Antal and Gronli,
2003; Demirbas, 2004).

Figure 2.1 Putative structure of charcoal (adopted from Bourke et al., 2007). A model of a
microcristalline graphitic structure is shown on on the left and an aromatic structure containing
oxygen and carbon free radicals on the right

It is commonly accepted that each biochar particle comprises of two main

structural fractions: stacked crystalline graphene sheets and randomly

51
ordered amorphous aromatic structures (Figure 2.1). Hydrogen, O, N, P and S
are found predominantly incorporated within the aromatic rings as
heteroatoms (Bourke et al., 2007). The presence of heteroatoms is thought to
be a great contribution to the highly heterogenous surface chemistry and
reactivity of biochar (see the next section).

2.1.2 Chemical composition and surface chemistry

Biochar composition is highly heterogeneous, containing both stable and
labile components (Sohi et al., 2009). Carbon, volatile matter, mineral matter
(ash) and moisture are generally regarded as its major constituents (Antal and
Gronli, 2003). Table 2.1 summarizes their relative proportion ranges in
biochar as commonly found for a variety of source materials and pyrolysis
conditions (Antal and Gronli, 2003; Brown, 2009).

Table 2.1 Relative proportion range of the four main components of biochar (weight percentage)
as commonly found for a variety of source materials and pyrolysis conditions (adapted from
Brown, 2009; Antal and Gronli, 2003)

Component Proportion (w w-1)

Fixed carbon 50-90
Volatile matter (e.g. tars) 0-40
Moisture 1-15
Ash (mineral matter) 0.5-5

The relative proportion of biochar components determines the chemical and

physical behaviour and function of biochar as a whole (Brown, 2009), which in
turn determines its suitability for a site specific application, as well as transport
and fate in the environment (Downie, 2009). For example, coarser and more
resistant biochars are generated by pyrolysis of wood-based feedstocks
(Winsley, 2007). In contrast, biochars produced from crop residues (e.g. rye,
maize), manures and seaweed are generally finer and less robust (lower
mechanical strength). The latter are also nutrient-rich, and therefore, more
readily degradable by microbial communities in the environment (Sohi et al.,
2009). The ash content of biochar is dependent on the ash content of the
biomass feedstock. Grass, grain husks, straw residues and manures
generally produce biochar with high ash contents, in contrast to that from
woody feedstocks (Demirbas 2004). For instance, manure (e.g. chicken litter)
biochars can contain 45% (by weight) as ash (Amonette and Joseph, 2009).
Moisture is another critical component of biochar (Antal and Gronli, 2003), as
higher moisture contents increase the costs of biochar production and
transportation for unit of biochar produced. Keeping the moisture content up
to 10% (by weight) appears to be desirable (Collison et al., 2009). In order for
this to be achieved, pre-drying the biomass feedstock may be a necessity,
which can be a challange in biochar production.

Despite the feasibility of biochar being produced from a wide range of

feedstocks under different pyrolysis conditions, its high carbon content and
strongly aromatic structure are constant features (Sohi et al., 2009).
According to Sohi et al. (2009), these features largely account for its chemical
stability. Similarly, pH shows little variability between biochars, and is tipically

52
>7. Table 2.2 summarizes total elemental composition (C, N, C:N, P, K,
available P – Pa - and mineral N) and pH ranges of biochars from a variety of
feedstocks (wood, green wastes, crop residues, sewage sludge, litter, nut
shells) and pyrolysis conditions (350-500oC) used in various studies (adapted
from Brown, 2009).

Table 2.2 Summary of total elemental composition (C, N, C:N, P, K, available P and mineral N)
and pH ranges and means of biochars from a variety of feedstocks (wood, green wastes, crop
residues, sewage sludge, litter, nut shells) and pyrolysis conditions (350-500ºC) used in various
studies (adapted from Chan and Xu, 2009)

pH C N N (NO3- C:N P Pa K
(g kg-1) (g kg-1) +NH4+) (g kg-1) (g kg-1) (g kg-1)
(mg kg-1)
Range From 6.2 172 1.7 0.0 7 0.2 0.015 1.0
To 9.6 905 78.2 2.0 500 73.0 11.6 58
Mean 8.1 543 22.3 - 61 23.7 - 24.3

Total carbon content in biochar was found to range between 172 to 905 g kg-
1
, although OC often accounts for < 500 g kg-1, as reviewed by Chan and Xu
(2009) for a variety of source materials. Total N varied between 1.8 and 56.4
g kg-1, depending on the feedstock (Chan and Xu, 2009). Despite seemingly
high, biochar total N content may not be necessarily beneficial to crops, since
N is mostly present in an unavailable form (mineral N contents < 2 mg k-1;
Chan and Xu, 2009). Nuclear magnetic resonance (NMR) spectroscopy has
shown that aromatic and heterocyclic N-containing structures in biochar occur
as a result of biomass heating, converting labile structures into more
recalcitrant forms (Almendros et al., 2003). C:N (carbon to nitrogen) ratio in
biochar has been found to vary widely between 7 and 500 Chan and Xu,
2009), with implications for nutrient retention in soils (see Sections 3.2.3). C:N
ratio has been commonly used as an indicator of the capacity of organic
substrates to release inorganic N when incorporated into soils.

Total P and total K in biochar were found to range broadly according to

feedstock, with values between 2.7 - 480 and 1.0 - 58.0 g kg-1, respectively
(Chan and Xu, 2009). Interestingly, total ranges of N, P and K in biochar are
wider than those reported in the literature for typical organic fertilizers. Most
minerals within the ash fraction of biochar are thought to occur as discrete
associations independent of the carbon matrix, with the exception of K and Ca
(Amonette and Joseph, 2009). Typically, each mineral association comprises
more than one type of mineral. Joseph et al. (2009) emphasize that our
current understanding of the role of high-mineral ash biochars is yet limited,
as we face the lack of available data on their long-term effect on soil
properties.

The complex and heterogeneous chemical composition of biochars is

extended to its surface chemistry, which in turn explains the way biochar
interacts with a wide range of organic and inorganic compounds in the
environment. Breaking and rearrangement of the chemical bounds in the
biomass during processing results in the formation of numerous functional
groups (e.g. hydroxyl -OH, amino-NH2, ketone -OR, ester -(C=O)OR, nitro -

53
NO2, aldehyde -(C=O)H, carboxyl -(C=O)OH) occurring predominantly on the
outer surface of the graphene sheets (e.g. Harris, 1997; Harris and Tsang,
1997) and surfaces of pores (van Zwieten et al., 2009). Some of these groups
act as electron donors, while others as electron acceptors, resulting on
coexisting areas which properties can range from acidic to basic and from
hydrophilic to hydrophobic (Amonette and Joseph 2009). Some functional
groups also contain other elements, such as N and S, particularly in biochars
from manures, sewage sludge and rendering wastes.

There is experimental evidence that demonstrates that the composition,

distribution, relative proportion and reactivity of functional groups within
biochar are dependent on a variety factors, including the source material and
the pyrolysis methodology used (Antal and Gronli, 2003). Different processing
conditions (temperature of 700oC or 450oC) explained differences in N
contents between three biochars from poultry litter (Lima and Marshall, 2005;
Chan et al., 2007). As the pyrolysis temperature rises, so does the proportion
of aromatic carbon in the biochar, while N contents peak at around 300oC
(Baldock and Smernik, 2002). In contrast, low processing temperatures
(<500oC) favour the relative accumulation of a large proportion of available K,
Cl (Yu et al., 2005), Si, Mg, P and S (Bourke et al., 2007; Schnitzer et al.,
2007). Therefore, processing temperatures < 500oC favour nutrient retention
in biochar (Chan and Xu, 2009) , while being equally advantageous in respect
to yield (Gaskin et al., 2008). Nevertheless, it is important to stress that
different permutations of those processing conditions, including temperature,
may affect differently each source material.

This emphasises the need for a case-by-case assessment of the chemical

and physical properties of biochar prior to its application into soil. Relating the
adverse effect of a particular constituent (or its concentration) of biochar to a
desirable biochar application rate (biochar loading capacity concept; Section
1.5.1) is difficult, as the exact biochar composition is often not provided in the
literature. The review of relevant literature has indicated that the full
knowledge on the composition of biochar as a soil amendment, and the way it
is influenced by those parameters, as well as the implications for soil
functioning, is still scarce. Partially, this can be explained by the fact that most
characterisation work has involved charcoals with high carbon and low ash
content, as required by the increasingly demanding market for activated
carbon. Another factor is the wide variety of processing conditions and
feedstocks available. The Black Carbon Steering Committee has developed
reference charcoal materials (from chestnut wood and rice grass) under
standardised pyrolysis conditions, representative of natural samples created
by forest fires, for comparison of quantification methods for BCs in soils and
sediments. Nevertherless, the current sparsity of biochar standards is largely
reflected on the poor understanding of the link between biochar composition
and its behaviour and function in soil.

2.2 Particle size distribution

Initially, particle size distribution in biochar is influenced mainly by the nature
of the biomass feedstock and the pyrolysis conditions (Cetin et al., 2004).

54
Shrinkage and attrition of the organic material occur during processing,
thereby generating a range of particle sizes of the final product. The intensity
of such processes is dependent on the pyrolysis technology (Cetin et al.,
2004). The implications of biochar particle size distribution on soils will be
discussed further throughout Chapter 3.

Particle size distribution in biochar also has implications for determining the
suitability of each biochar product for a specific application (Downie et al.,
2009), as well as for the choice of the most adequate application method (see
Section 1.8). In addition, health and safety issues relating to handling, storage
and transport of biochar are also largely determined by its particle size
distribution, as discussed in this report in regard to its dust fraction (see
Sections 2.2.1 and 5.2).

The influence of the type of feedstock on particle size distribution was

discussed by Sohi et al. (2009), among others. Wood-based feedstocks
generate biochars that are coarser and predominantly xylemic in nature,
whereas biochars from crop residues (e.g. rye, or maize) and manures offer a
finer and more brittle structure (Sohi et al., 2009). Downie et al. (2009) have
further provided evidence of the influence of feedstock and processing
conditions on particle size distribution in biochar. Sawdust and woodchips
under different pre-treatments were pyrolised using continuous slow pyrolysis
(heating rate of 5-10ºC min-1), after which particle size distribution in the
resulting biochar was assessed through dry sieving. Generally, particle size
was found to decrease as the pyrolysis heat treatment temperature increased
(450ºC-700ºC range) for both feedstocks, due to a reduction of the biomass
material resistance to attrition during processing (Downie et al., 2009).

The operating conditions during pyrolysis (e.g. heating rate, high treatment
temperature -HTT, residence time, pressure, flow rate of the inert gas, reactor
type and shape) and pre- (e.g. drying, chemical activation) and post- (e.g.
sieving, activation) treatments can greatly affect biochar physical structure
(Gonzalez et al., 1997; Antal and Grønli, 2003; Cetin et al., 2004; Lua et al.,
2004; Zhang et al., 2004; Brown et al., 2006). Such observations were derived
mainly from studies involving activated carbon produced from a variety of
feedstocks, including maize hulls (Zhang et al., 2004), nut shells (Lua et al.,
2004; Gonzaléz et al., 2009) and olive stones (Gonzaléz et al., 2009).
Similarly, heating rate, residence time and pressure during processing were
shown to be determinant factors for the generation of finer biochar particles,
independently of the original material (Cetin et al., 2004). For instance, for
higher heating rates (e.g. up to 105-500ºC sec-1) and shorter residence times,
finer feedstock particles (50-2000 µm) are required in order to facilitate heat
and mass transfer reactions, resulting in finer biochar material (Cetin et al.,
2004). In contrast, slow pyrolysis (heating rates of 5-30ºC min-1) can use
larger feedstock particles, thereby producing coarser biochars (Downie et al.,
2009). Increasing the proportion of larger biochar particles can also be
obtained by increasing the pressure (from atmospheric to 5, 10 and 20 bars)
during processing, which was explained by both particle swelling and
clustering, as a result of melting (i.e. plastic deformation) followed by fusion
(Cetin et al., 2004).

55
2.2.1 Biochar dust
The term ‘dust’ is described in this report as referring to the fine and ultrafine
fraction of biochar, comprising various organic and inorganic compounds of
distinct particle sizes within the micro- and nano-size range (Harris and
Tsang, 1997; Cornelissen et al., 2005). Harris and Tsang (1997) researched
the micro- and nano-sized fraction of chars, although so far, this issue
remains poorly understood. Biomass precursor (feedstock) and the pyrolysis
conditions (Donaldson et al., 2005; Hays and van der Wal, 2007) are likely to
be primary factors influencing the properties of biochar dust (Downie et al.,
2009), including the type and size of its particles, as well as the proportion of
micro- and nanoparticles, as discussed previously
Harris and Tsang (1997) used high resolution electron microscopy (HREM) for
studying the smaller fraction of charcoal resulting from the pyrolysis (700ºC) of
sucrose and concluded that charcoal dust consists of round fullerene-like
nanoparticles (Harris and Tsang, 1997). Brodowski et al. (2005) corroborates
the finding of porous spherical-shaped particles (with surface texture ranging
from smooth to rough) within the <2 µm fraction of charcoals in a field-plot
topsoil (0-10 cm), although no reference to the word “fullerene” was found.
What is important in this context is that, considering the small size of such
particles and their reactivity, the proportion of dust within the biochar (which
may also apply to biochars with high ash contents) has relevant practical, as
well as health and safety implications (see Section 5.2).
The proportion of dust in biochar is also key in determining the suitability of a
given application strategy (Blackwell et al., 2009). For example, Holownicki
(2000) suggested that this fine fraction could be successfully employed in
precision agriculture for spraying fungicide preparations in orchards and
vineyards. When injection is appropriate, Blackwell et al. (2009) pointed out
that the application of biochar dust may in fact be preferred when used in
combination with liquid manure in selected crops.
On the other hand, biochar dust has been identified in the literature as a
better sorbent for a wide range of trace hydrophobic contaminants (e.g. PAHs,
polychlorinated biphenyls - PCBs, pesticides, polychlorinated dibenzeno-p-
dioxins and –furans - PCDD/PCDFs), when compared to larger biochar
particles or to particulate organic matter (Hiller et al., 2007; Bucheli and
Gustafsson, 2001, 2003). As such, the addition of biochar dust to soils may
increase the sorption affinity of the soil for common environmental pollutants
(see Section 3.2.2 for a more detailed discussion on the sorption of
hydrophobic compounds to biochar), as demonstrated for dioxin sorption in a
marine system (Persson et al., 2002).

2.3 Pore size distribution and connectivity

Biomass feedstock and the processing conditions are the main factors
determining pore size distribution in biochar, and therefore its total surface
area (Downie et al., 2009). During thermal decomposition of biomass, mass
loss occurs mostly in the form of organic volatiles, leaving behind voids, which
form an extensive pore network. This section focuses on pore size distribution

56
in biochar, while biochar density is discussed in the context of the biochar-soil
mixture in Section 3.1.1.

Biochar pores are classified in this review into three categories (Downie et al.,
2009), according to their internal diameters (ID): macropores (ID >50 nm),
mesopores (2 nm< ID <50 nm) and micropores (ID <2 nm). These categories
are orders of magnitude different to the standard categories for pore sizes in
soil science (see Table 3.1). The elementary porosity and structure of the
biomass feedstock is retained in the biochar product formed (Downie et al.,
2009). The vascular structure of the original plant material, for example, is
likely to contribute for the occurrence of macropores in biochar, as
demonstrated for activated carbon from coal and wood precursors (Wildman
and Derbyshire, 1991). In contrast, micropores are mainly formed during
processing of the parent material. While macropores have been were
identified as a ‘feeder’ to smaller pores (Martinez et al., 2006), micropores
effectively account for the characteristically large surface area in charcoals
(Brown, 2009).

Among those operating parameters, HTT is thought to be the most significant

factor for the resulting pore distribution in charcoals (Lua et al., 2004), as the
physical changes undergone by the biomass feedstock during processing are
often temperature-dependent (Antal and Grønli, 2003).

The development of microporosity in biochar, which is linked to an increase in

structural and organisational order, has been showed to be favoured by
higher HTT and retention times, as previously demonstrated for activated
carbon (e.g. Lua et al., 2004). For example, increasing pyrolysis temperature
from 250 to 500oC enhanced the development of micropores in chars derived
from pistachio-nut shells, due to increased evolution of volatiles. For
subsequent increases in temperature (>800oC), a reduction of the overall
surface area of the char was observed and was attributed to partial melting of
the char structure (Lua et al., 2004). Similarly, heating rate and pressure
during processing have also been found to influence the mass transfer of
volatiles produced at any given temperature range, and are therefore
regarded as key contributing parameters influencing pore size distribution
(Antal and Grønli, 2003). For instance, Lua et al. (2004) observed a peak in
surface area of pistachio-nut shell char at low heating rates (10oC), whereas
higher heating rates resulted in a decrease in surface area.

It is important to stress, however, that the relative influence of each

processing parameter on the final microporosity in biochar is determined by
the type of feedstock, as noted from the above studies (e.g. Cetin et al., 2004;
Lua et al., 2004; Pastor-Villegas et al., 2006; Gonzaléz et al., 2009). In
particular, the lignocellulosic composition of the parent material largely
determines the rate of its thermal decomposition, and therefore, the
development of porosity (Gonzaléz et al., 2009). In the case of charcoals from
almond tree pruning, a greater volume of meso and macropores was
obtained, which was accounted for by the slow decomposition rate of such
precursor during the initial stages of pyrolysis (Gonzaléz et al., 2009). The

57
opposite was found for almond shell, probably due to its inherently high initial
thermal decomposition rate (Gonzaléz et al., 2009).

2.4 Thermodynamic stability

The thermodynamic equilibrium concerning carbonised residues, such as
biochar, favours the production of CO2.

C ( graphite ) + O2 (g ) → CO2
( )
Equation 1
ΔH o f = −393.51 kJ .mol −1
The standard enthalpy of formation is represented as ΔH°f.and the degree sign denotes the
standard conditions (P = 1 bar and T = 25°C)

Equation 1 shows that the oxidation of graphite, being the most

thermodynamically stable form of carbon, will occur spontaneously as shown
by the negative energy value (meaning that 393.51 kJ of energy is emitted for
every mole of CO2 ‘produced’). Since the oxidation of graphite to carbon
dioxide will occur, allbeit very slowly under normal conditions (Shneour,
1966), all other forms of carbon which are less thermodynamically stable than
graphite, will also undergo oxidation to CO2 in the presence of oxygen. The
speed at which this oxidation occurs depends on a number of factors, such as
the precise chemical composition, as well as the temperature and moisture
regime to which the compound is exposed. Furthermore, residence time of
biochar in soils will also be affected by microbial processes. The recalcitrance
of biochar in soil is discussed in more depth in Sections 3.2.1 and 3.2.5.1.

2.5 CEC and pH

CEC variation in biochars ranges from negligible to around 40 cmolc g-1 and
has been reported to change following incorporation into soils (Lehmann,
2007). This may occur by a process of leaching of hydrophobic compounds
from the biochar (Briggs et al., 2005) or by increasing carboxylation of C via
abiotic oxidation (Cheng et al. 2006; Liang et al. 2006). Glaser et al. (2001)
discussed the importance of ageing to obtain the increases in CEC of black
BC found in the Terra Preta soils of the Amazon.
Considering the very large heterogeneity of its properties, biochar pH values
are relatively homogeneous, that is to say they are largely neutral to basic.
Chan and Xu (2009) reviewed biochar pH values from a wide variety of
feedstocks and found a mean of pH 8.1 in a total range of pH 6.2 – 9.6. The
lower end of this range seems to be from green waste and tree bark
feedstocks, with the higher end from poultry litter feedstocks.

2.6 Summary
Biochar is comprised of stable carbon compounds created when biomass is
heated to temperatures between 300 to 1000°C under low (preferably zero)
oxygen concentrations. The structural and chemical composition of biochar is
highly heterogeneous, with the exception of pH, which is tipically > 7. Some
properties are pervasive throughout all biochars, including the high C content

58
and degree of aromaticity, partially explining the high levels of biochar’s
inherent recalcitrance. Neverthless, the exact structural and chemical
composition, including surface chemistry, is dependent on a combination of
the feedstock type and the pyrolysis conditions (mainly temperature) used.
These same parameters are key in determining particle size and pore size
(macro, meso and micropore; distribution in biochar. Biochar's physical and
chemical characteristics may significantly alter key soil physical properties
and processes and are, therefore, important to consider prior to its application
to soil. Furthermore, these will determine the suitability of each biochar for a
given application, as well as define its behaviour, transport and fate in the
environment. Dissimilarities in properties between different biochar products
emphasises the need for a case-by-case evaluation of each biochar product
prior to its incorporation into soil at a specific site. Further research aiming to
fully evaluate the extent and implications of biochar particle and pore size
distribution on soil processes and functioning is essential, as well as its
influence on biochar mobility and fate (see Section 3.2.1).

59
3. EFFECTS ON SOIL PROPERTIES, PROCESSES
AND FUNCTIONS
This chapter discusses the effects of biochars with different characteristics
(Chapter 2) on soil properties and processes. First, effects on the soil
properties are discussed, followed by effects on soil physical, chemical and
biological processes. The agricultural aspect of the production function of soil
is reviewed in detail (including meta-analyses)

3.1 Properties
3.1.1 Soil Structure
The incorporation of biochar into soil can alter soil physical properties such as
texture, structure, pore size distribution and density with implications for soil
aeration, water holding capacity, plant growth and soil workability (Downie et
al., 2009). Particularly in relation to soil water retention, Sohi et al. (2009)
propose an analogy between the impact of biochar addition and the observed
increase in soil water repellency as a result of fire. Rearrangement of
amphiphilic molecules by heat from a fire, as proposed by Doerr et al. (2000),
would not affect the soil, but could affect the biochar itself during pyrolysis. In
addition, the soil hydrology may be affected by partial or total blockage of soil
pores by the smallest particle size fraction of biochar, thereby decreasing
water infiltration rates (see Sections 3.1.1 and 3.2.3). In that sense, further
research aiming to fully evaluate the extent and implications of biochar
particle size distribution on soil processes and functioning is essential, as well
as its influence on biochar mobility and fate (see Section 3.2.1).

3.1.1.1 Soil Density

Biochar has a bulk density much lower than that of mineral soils and,
therefore, application of biochar can reduce the overall bulk density of the soil,
although increases in bulk density are also possible. If 100 t ha-1 of biochar
with a bulk density of 0.4 g cm-3 is applied to the top 20 cm of a soil with a
bulk density of 1.3 g cm-3, and the biochar particles do not fill up existing soil
pore space, then the soil surface in that field will be raised by ca. 2.5 cm with
an overall bulk density reduction (assuming homogeneous mixing) of 0.1 g
cm-3 to 1.2 g cm-3. However, if the biochar that is applied has a low
mechanical strength and disintegrates relatively quickly into small particles
that fill up existing pore spaces in the soil, then the dry bulk density of the soil
will increase.
In agronomy, relatively small differences in soil bulk density can be associated
with agronomic benefits. Conventionally, i.e. without biochar additions, lower
bulk density is associated with higher SOM content leading to nutrient release
and retention (fertiliser saving) and/or lower soil compaction due to better soil
management (potentially leading to improved seed germination and cost
savings for tillage and cultivation). Biochar application to soil by itself may
improve nutrient retention directly (see Section 3.2.2), but nutrient release is
mostly very small (except for some biochars in the first years, especially in
ash-rich biochars) and the application of biochar with heavy machinery may
compact the subsoil, depending on the application method and timing

61
Soil compactibility is closely related to soil bulk density. Soane (1990)
reviewed the effect of SOM, i.e. not including biochar, on compactibility and
proposed several mechanisms by which SOM may influence the ability of the
soil to resist compactive loads:
1) Binding forces between particles and within aggregates. Many of the
long-chain molecules present in SOM are very effective in binding
mineral particles. This is of great importance within aggregates which
“…are bound by a matrix of humic material and mucilages” (Oades in
Soane, 1990).
2) Elasticity. Organic materials show a higher degree of elasticity under
compression than do mineral particles. The relaxation ratio – R – is
defined as the ratio of the bulk density of the test material under
specified stress to the bulk density after the stress has been removed.
Relaxation effects of materials such as straw are therefore much
greater than material like slurry or biochar.
3) Dilution effect. The bulk density of SOM is usually appreciably lower
than mineral soil. It can however differ greatly, from 0.02 t m-3 for some
types of peat to 1.4 t m-3 for peat moss, compared to 2.65 t m-3 for
mineral particles (Ohu et al. in Soane, 1990).
4) Filament effect. Roots, fungal hyphae and other biological filaments
have the capacity to bind the soil matrix.
5) Effect on electrical charge. Solutions/suspensions of organic
compounds may increase the hydraulic conductivity of clays by
changing the electrical charge on the clay particles causing them to
move closer together, flocculate and shrink, resulting in cracks and
increased secondary – macro - porosity (Soane, 1990). Biochar’s ash
fraction could cause similar effects.
6) Effect on friction. An organic coating on particles and organic material
between particles is likely to increase the friction between particles
(Beekman in: Soane, 1990). The direct effect of biochar on soil friction
has not been studied.
The effect of biochar application on soil compactibility has not been tested
experimentally yet. From the above mechanisms, however, direct effects of
biochar are probably mostly related to bullet points 3, 5 and 6 above. The very
low elasticity of biochar suggests that resilience to compaction, i.e. how
quickly the soil ‘bounces back’, is unlikely to be increased directly by biochar.
The resistance to compaction of soil with biochar could potentially be
enhanced via direct or indirect effects (interaction with SOM dynamics and
soil hydrology). For example, some studies have shown an increase in
mycchorizal growth after additons of biochar to soil (see Section 3.2.6) while
under specific conditions plant productivity has also been shown to increase
(see Section 3.3). The enhanced development of hyphae and roots will have
an effect on soil compaction However, experimental research into the
mechanisms and subsequent modeling work is required before any
conclusions can be drawn regarding the overall effect of biochar on soil
compaction.

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3.1.1.2 Soil pore size distribution
The incorporation of biochar into soil can alter soil physical properties such as
texture, structure, pore size distribution and density with implications for soil
aeration, water holding capacity, plant growth and soil workability. The soil
pore network can be affected by biochar’s inherent porosity as well as its
other characteristics, in several ways. Biochar particle size and pore size
distribution and connectivity, the mechanical strength of the biochar particles,
and the translocation and interaction of biochar particles in the soil are all
determining factors that will lead to different outcomes in different soil-climate-
management combinations. As described in the above section, these factors
can cause the overall porosity of the soil to increase or decrease following
biochar incorporation into soils.

There is evidence that suggests that biochar application into soil may increase
the overall net soil surface area (Chan et al., 2007) and consequently, may
improve soil water retention (Downie et al., 2009; see Section 3.1.2) and soil
aeration (particularly in fine-textured soils; Kolb, 2007). An increased soil-
specific surface area may also benefit native microbial communities (Section
3.2.6) and the overall sorption capacity of soils (Section 3.2.2). In addition, soil
hydrology may be affected by partial or total blockage of soil pores by the
smallest particle size fraction of biochar, thereby decreasing water infiltration
rates (see Sections 3.1.1, 3.1.2 and 3.2.3). Nevertheless, experimental
evidence of such mechanisms is scarce and, therefore, any effects of the pore
size distribution of biochar on soil properties and functions is still uncertain at
this stage. Further research aiming to fully evaluate the extent and
implications of biochar particle size distribution on soil processes and
functioning is essential, as well as its influence on biochar mobility and fate in
the environment (see Section 3.2.1).

Table 3.1 shows the classifications of pore sizes in material science and soil
science. Fundamental differences, i.e. orders of magnitude difference for
classes with the same names, are obstacles in communicating to any
audience outside of biochar research and also hinder the communication
efficiency within interdisciplinary research groups that work on biochar in soils.
Therefore, it is recommended that existing classifications are modified to
resolve this confusion. However, in this review we will use the existing
terminology and the relevant classification will need to be retrieved from the
context.

Table 3.1 Pore size classes in material science vs. soil science

Material science Soil science

Pore size (µm)
Cryptospores na <0.1
Ultramicropores na 0.1-5
Micropores <0.002 5-30
Mesopores 0.002-0.05 30-75
Macropores >0.05 >75

63
3.1.2 Water and Nutrient Retention
The addition of biochar to soil will alter both the soil’s chemical and physical
properties. The net effect on the soil physical properties will depend on the
interaction of the biochar with the physicochemical characteristics of the soil,
and other determinant factors such as the climatic conditions prevalent at the
site, and the management of biochar application.
Adding biochar affects the regulation and production function of the
agricultural soil. To what extent biochar is beneficial to agriculture, and the
dominant mechanisms that determine this, is still under scientific scrutiny.
Agronomic benefits of biochar are often attributed to improved water and/or
nutrient retention. However, many of the scientific studies are limited to site-
specific soil conditions, and performed with biochar derived from specific
feedstocks. Of more concern, and as of yet underexposed, is the stability of
the structural integrity of the biochar. Especially when biochar is used in
today’s intensive agriculture with the use of heavy machinery, opposed to the
smallholder system that led to the formation of Terra Preta. Another concern
relates to the potential externalities of bringing large quantifies of biochar in
the environment (see Chapter 5).
The mechanisms that lead to biochar-provided potential improvements in
water retention are relatively straightforward. Adding biochar to soil can have
direct and indirect effects on soil water retention, which can be short or long
lived. Water retention of soil is determined by the distribution and connectivity
of pores in the soil-medium, which is largely regulated by soil particle size
(texture), combined with structural characteristics (aggregation) and SOM
content.
The direct effect of biochar application is related to the large inner surface
area of biochar. Biochars with a range in porous structures will result from
feedstocks as variable as straw, wood and manure (see Sections 1.7, 2.1 and
2.3). Kishimoto and Sugiura (1985) estimated the inner surface area of
charcoal formed between 400 and 1000°C to range from 200 to 400 m2 g-1.
Van Zwieten et al. (2009) measured the surface area of biochar derived from
papermill waste with slow pyrolysis at 115 m2 g-1.
The hypothesised indirect effects of biochar application on water retention of
soil relate to improved aggregation or structure. Biochar can affect soil
aggregation due to interactions with SOM, minerals and microorganisms. The
surface charge characteristics, and their development over time, will
determine the long term effect on soil aggregation. Aged biochar generally
has a high CEC, increasing its potential to act as a binding agent of organic
matter and minerals. Macro-aggregate stability was reported to increase with
20 to 130% with application rates of coal derived humic acids between 1.5 Mg
ha-1 and 200 t ha-1 (Mbagwu and Piccolo, 1997). Brodowski et al (2006) found
indications that BC acted as a binding agent in microaggregates in soils under
forest, grassland and arable land use in Germany. In-situ enhancement of soil
aggregation by biochar requires further analysis.
The mechanical stability and recalcitrance of biochar once incorporated in the
soil will determine long term effects on water retention and soil structure. This
is determined by feedstock type and operating conditions as well as the
prevalent physical-chemical conditions that determine its weathering and the

64
compaction and compression of the biochar material in time. The effect of the
use of heavy agricultural machinery on compaction of the soil-biochar matrix
has yet to be studied in detail. Another factor contributing to the uncertainty in
long-term beneficial effects of biochar application to soil is the potential
clogging or cementation of soil pores with disintegrated biochar material.
Glaser et al. (2002b) reported that Anthrosols rich in charcoal with surface
areas three times higher than those of surrounding soils had an increased
field capacity of 18%. Tryon (1948) studied the effect of charcoal on the
percentage of available moisture in soils of different textures. In sandy soil the
addition of charcoal increased the available moisture by 18% after adding
45% of biochar by volume, while no changes were observed in loamy soil,
and in clayey soil the available soil moisture decreased with increasing coal
additions. This was attributed to hydrophobicity of the charcoal, although
another factor could simply be that the biochar was replacing clay with a
higher water retention capacity. Biochar’s high surface area can thus lead to
increased water retention, although the effect seems to depend on the initial
texture of the soil. Therefore, improvements of soil water retention by charcoal
additions may only be expected in coarse-textured soils or soils with large
amounts of macropores. A draw-back is the large volume of biochar that
needs to be added to the soil before it leads to increased water retention.
The capacity of the agricultural soil to store water regulates the time and
amount water is kept available for crop transpiration. Tseng and Tseng (2006)
found that activated biochar contained over 95% of micropores with a
diameter <2 nm. Since the porosity of biochar largely consists of micropores,
the actual amount of additional plant available water will depend on the
biochar feedstock and the texture of the soil it is applied to. The agronomic
water-storage benefit of biochar application will thus dependent on the relative
modification of the proportion of micro, meso and macro pores in the root
zone. In sandy soils, the additional volume of water and soluble nutrients
stored in the biochar micropores may become available as the soil dries and
the matric potential increases. This may lead to increased plant water
availability during dry periods.
The potential co-benefits or negative externalities of the use of biochar in
irrigated agricultural systems have not been explored in detail. If the water
holding capacity of the soil increases this may hypothetically reduce the
irrigation frequency or irrigation volume. However, the potential susceptibility
of disintegrated biochar particles to cement or clog the soil may also result in
increased runoff and lower infiltration rates.

65
 0.70 Standard soil (van
Genuchten, 1980)

Plus biochar?
0.60

0.50
Soil water content [cm3/cm3]

0.40

0.30

0.20

0.10

0.00
0.1 1 10 100 1000 10000 100000 1000000
Pressure head [-kPa]

Figure 3.1 Typical representation of the soil water retention curve as provided by van Genuchten
(1980) and the hypothesized effect of the addition of biochar to this soil

Figure 3.1 shows a typical representation of the soil water retention curve
(van Genuchten, 1980) and the hypothesised effect of the addition of biochar
to this soil. Notice that in this conceptual example most of the water that is
stored additionally in the soil will not be available for plant water uptake since
it occurs at tensions superior to the range wherein plant roots are able to take
up water. In this hypothetical representation this is mainly due to the pore size
distribution of the biochar which largely consists of very small pores and only
very little pores in the range relevant for plant water uptake. Although this is a
hypothetical consideration; it highlights the need for a further understanding of
the direct and indirect effects of biochar addition on soil water retention, and
its longevity.

3.1.2.1 Soil water repellency

Soil water repellency (SWR), or hydrophobicity, is defined functionally as “the
reduction of the affinity of soils to water such that they resist wetting for
periods ranging from a few seconds to hours, days or weeks” (King, 1981).
SWR is a widespread phenomenon associated with decreased infiltration
rates, fingered flow infiltration, and increased runoff. In the case of agricultural
land, fertiliser and biocide (herbicide, pesticide) leaching to the groundwater
via bypass flow (secondary porosity) can be costly to the farmer and the

66
environment. Most of the literature on soil water repellency focuses the effect
of the heat wave from a (wild)fire on the hydrophobic properties of the SOM.
Reorientation of amphiphilic molecules is one of the hypothesised
mechanisms (Doerr et al., 2000) explaining the water repellent effect,
although other mechanisms are also hypothesised. In relation to soil water
retention, Sohi et al. (2009) propose an analogy between the impact of
biochar addition and the observed increase in soil water repellency as a result
of fire. Rearrangement of amphiphilic molecules by heat from a fire, as
proposed by Doerr et al. (2000), would not affect the soil, but could affect the
biochar itself during pyrolysis.
Field studies on water repellent properties of biochar or charcoal are absent
from the scientific literature and very limited even for charcoal produced by
wildfires. Briggs et al. (2005) measured WR of charcoal particles after a
wildfire in a pine forest and found very large differences in WR between
charcoal particles on the surface and in the mineral soil vs. those on the
border of the litter layer and mineral soil. The water drop penetration time, that
is the time it takes a droplet of water to infiltrate, was >2 h for the former and
<10 s for the latter. The authors proposed leaching by organic acids as a
mechanism explaining the reduction of water repellent properties underneath
the litter layer. How biochar may influence soil water repellency, directly or
indirectly, is a topic that still requires a substantial research effort before the
mechanisms are understood and predictions can be made. A trade off
appears to exist between the capacity to bind HOCs, like PAHs (see Section
3.2.2), and the capacity to bind water molecules.

3.1.3 Soil colour, albedo and warming

From the Anthrosol profile pictures (Figure 1.5) it is obvious that high
concentrations of biochar in soil darken its colour. Briggs et al. (2005)
measured changes in dry soil colour from charcoal additions and found the
Munsell value to decrease from 5.5 to 4.8 at charcoal concentrations of 10 g
kg-1, and down to 3.6 at 50 g kg-1. Oguntunde et al. (2008) compared the soil
colour of charcoal sites (i.e. where charcoal used to be produced) with that of
adjacent soil and found the Munsell value to decrease from 3.1 (± 0.6) to 2.5
(± 0.4). The degree of darkening is dependent on i) the colour of the soil prior
to biochar additions (Munsell value 1-9), ii) the colour of the biochar (probably
Munsell value 0-2), iii) the biochar concentration in the soil, iv) the degree of
mixing (related to particle size of both the biochar and the soil), v) the surface
roughness, and vi) the change in water retention at the soil surface that
accompanies the addition of biochar (moist soil is darker in colour). Wang et
al. (2005) conducted three years of continuous measurement in a semi-desert
area in Tibet and showed an exponential relationship between soil moisture
content (v v-1) and surface albedo. The combined effects of the changes in
these factors subsequently determine the albedo effect of the soil.

Land surface albedo is an important component of global and regional climate

change models. However, almost exclusively, the albedo of the vegetation is
used, not that of soil. Levis et al. (2004) introduced a modification to soil
albedo into their community climate system model and found this change to

67
be the key for the model output to resemble the botanic evidence for climate-
vegetation interactions in mid-Holocene North Africa. Model simulations with a
darker soil colour led to an intensified monsoon which brought precipitation
further north; testifying the importance of changes in soil albedo on climate
feedbacks.

The principle that biochar application to soils decreases the albedo of bare
soil and thereby contributes to further warming of the planet is accepted,
however, if, and where, that would lead to an effect of relevant magnitude is
much less certain. Bare soil is limited to the winter months on fields growing
spring crops, or in orchards without ground cover (e.g. olive orchards,
vineyards). In the former case, the warming effect may be relatively small
because solar radiation reaching the surface is low in winter months,
however, many orchards and vineyards are in more southern parts that
receive a greater solar input and the bare soil conditions persist throughout
the year. Post et al. (2000) investigated the influence of soil colour and
moisture content on the albedo of 26 different soils ranging widely in colour
and texture. They found that wet samples had their albedo reduced by a
mean of 48% (ranging between 32-58%), and that Munsell colour value is
linearly related to soil albedo.
The amount of solar radiation that reaches the soil surface (as affected by sun
angle and slope and vegetation cover) and the specific heat of soils, largely
control the rate at which soils warm up in the spring, and thus influence the
emergence of seedlings. Soil colour and soil moisture content are the main
factors determining the specific heat of soil. For pure water the specific heat is
about 4.18 J g-1 K-1; that of dry soil is about 0.8 J g-1 K-1. Therefore, although
soils high in biochar content are usually dark in colour, if the biochar increases
the water retention of the soil concomitantly (see Section 3.1.2) then the
associated extra energy absorption is countered by a high water content,
which causes the soil to warm up much more slowly (Brady, 1990). This
implies that biochar with low water retention capacity (e.g. because of water
repellent properties, see Section 3.1.2.1) will cause the greatest increase in
soil warming, and that this impact will be greatest where biochar is applied to
light-coloured soils (high Munsell value) with spring crops (i.e. bare soil in
spring) or orchards/vineyards.

3.1.4 CEC and pH

The cation exchange capacity (CEC) of soils is a measure for how well some
nutrients (cations) are bound to the soil, and, therefore, available for plants
uptake and ‘prevented’ from leaching to ground and surface waters. It is at
negatively charged sites on the reactive surface area of biochar (and clay and
organic matter) where cations can be electro-statically bound and exchanged.
Cations compete with each other as well as with water molecules and can be
excluded when the pore size at the charged site is smaller than their size.
Cheng et al. (2006) assessed the effects of climatic factors on biochar
oxidation in natural systems. The CEC of biochar was correlated to the mean
temperature and the extent of biochar oxidation was related to its external
surface area, being seven times higher on the external surfaces than in its
interior (Cheng et al., 2008). It is not known at present how the CEC of

68
biochar will change as the biochar disintegrates by weathering and tillage
operations, ‘ages’ and moves through the soil.
Anions are bound very poorly by soils under neutral or basic pH conditions.
This is one of the reasons why crops need fertilising, as anionic nutrients (e.g.
phosphates) are leached or flushed from the soil into ground/surface waters
(eutrophication). Cheng et al. (2007) found biochar to exhibit an anion
exchange capacity (at pH 3.5) which decreased to zero as it aged in soil (over
70 years). If biochar can play a role in anion exchange capacity of soils
remains an unanswered question and a research effort is required into the
mechanisms to establish under what conditions (e.g. more neutral pH) anions
may be retained.
As previously discussed, biochar pH is mostly neutral to basic (see Table 2.2).
The liming effect has been discussed in the literature as one of the most likely
mechanisms behind increases in plant productivity after biochar applications,
and the meta-analysis in this report (Section 3.3) provides supporting
evidence for that mechanism. Lower pH values in soils (greater acidity) often
reduce the CEC and thereby the nutrient availability. In addition, for many of
the tropical soils studied, reduced aluminium toxicity by reducing the acidity is
proposed as the most likely chemical mechanism behind plant productivity
increases.
For the experimental studies used in the meta-analysis on plant productivity
(see Section 3.3.1) the average pre-amendment soil pH was 5.3 and post-
amendment 6.2, although for poultry litter biochar on acidic soils the change
was as large as from pH 4.8 to 7.8. Therefore, a scientific consensus on a
short term liming effect of biochar applied to soil is apparent. This implies that
biochars with greater liming capacity can provide greater benefit to arable
soils that require liming, by being applied more frequently at lower application
rates. Thereby reducing, or potentially cutting, a conventional liming
operation, and hence providing a clear cost saving.

3.2 Soil Processes

3.2.1 Environmental behaviour, mobility and fate
An effective evaluation of biochar stability in the environment is paramount,
particularly when considering its feasibility as a carbon sequestration tool. A
sound understanding of the contribution that biochar can make to improve soil
processes and functioning relies on knowing the extent and implications of the
changes biochar undergoes in soil over time. Such knowledge remains,
however, sparse and most experimental evidence has been gathered for
other forms of black carbon. Energy-dispersive X-ray spectrometry looks
promising as a tool for providing evidence of such changes in soil (Glaser et
al., 2000; Brodowski et al., 2005a).
Current evaluations of the age of black carbon particles from both wildfires
and anthropogenic activity indicate great stability of (at least) a significant
component of biochar, ranging from several millennia to hundreds of years
(e.g. Skjemstad et al., 2001; Lehmann et al., 2009). Such stability has been
employed as a tool for evaluating, dating and modelling of ancient cropping

69
and management practices (Scott et al., 2000; Ferrio et al., 2006). Yet,
establishing the mean residence time of biochars in natural systems remains
a challenge, partly due to their inherent heterogeneity, and partly due to
different interactions with both the biotic (e.g. microbial communities, flora)
and abiotic (e.g. clays, humic substances) components of soil (Brodowski et
al., 2005a, 2006).
Analysis of biochar-enriched agricultural soil using X-ray spectrometry and
scanning electron microscopy showed that biochar particles in soil occur
either as discrete particles or as particles embedded and bound to minerals
(mainly clay and silt; Brodowski et al., 2005). This corroborates earlier studies
reporting that most biochar in Amazonian Terra Preta was found in the light
(<0.2 g cm-3) fraction of soil (Gu et al., 1995), which Hammes and Schmidt
(2009) refer to as “intrinsically refractory”, while a minor amount occurred
adsorbed to the surface of mineral particles (Gu et al., 1995). It is also likely
that a significant portion of biochar occurs in aggregate-occluded organic
matter in soil (see Section 3.2.5.3).
Biochar is no longer considered inert, although mechanisms involved in
biochar degradation in soil not being fully understood (Hammes and Schmidt,
2009). It has been demonstrated that exposure to strong chemical oxidants
(e.g. Skjemstad et. al., 1996), including ozone (Kawamoto et al., 2005), and to
high temperatures (Morterra et al., 1984; Cheng et al., 2006) can cause
oxidation in charcoal over short periods of time. In natural environments,
photochemical and microbial breakdown appear to be the primary degradation
mechanisms (Goldberg, 1985), which can result in alteration of the charcoal’s
surface chemistry and functional properties (e.g. CEC, nutrient retention;
Glaser et al., 2002). Such mechanisms have been assessed by a relatively
small number of short-term experiments involving biochar-enriched soils in the
presence and absence of added substrates (e.g. Hamer et al., 2004; Cheng et
al., 2006). Incubation studies appear to indicate that biological decomposition
is very slow (see Section 3.2.5.1) and might be of minor relevance compared
to abiotic degradation (see Section 3.2.5.1), particularly when fresh biochars
are concerned (Cheng et al., 2006).
Surfaces of fresh biochars are generally hydrophobic and have relatively low
surface charges (Lehmann et al., 2005). However, over time, biochar
oxidation in the soil environment due to aging, may reflect in accumulation of
carboxylic functionalities at the surfaces of biochar particles (Brodowski et al.,
2005), promoting, perhaps, further interactions between biochar and other soil
components (Cheng et al., 2006), including organic and mineral matter
(Brodowski et al., 2005), as well as contaminants (Smernik et al., 2006). It is
reasonable to hypothesize that solubilisation, leaching and translocation of
biochar within the soil profile and into water systems is also expected to be
gradually enhanced for longer exposure periods in soil (Cheng et al., 2006).
Whether the relative importance of microbial decomposition increases over
time (as biochar particle size decreases) remains largely unknown and
attempts to determine actual mineralisation rates are still scarce.
Although biochar characteristics (e.g. particle and pore size distribution,
surface chemistry, relative proportion of readily available components), as
well as physical and chemical stabilisation mechanisms may contribute to the

70
long mean residence times of biochar in soil, the relative contribution of each
factor to short- and long-term biochar loss has been poorly assessed,
particularly when influenced by environmental conditions. Biochar
characteristics are largely determined by the feedstock and pyrolysis
conditions, as previously discussed. For instance, particle size is likely to
influence the rate of both abiotic and biotic degradation in soil, as
demonstrated for biochar particles >50 µm in a Kenyan Oxisol (Nguyen et al.,
2008 in Lehmann et al., 2009). Therefore, processes which favour biochar
fragmentation into smaller particles (e.g. freeze-thaw cycles, rain and wind
erosion, bioturbation) may not only enhance its degradation rate, but also
render it more susceptible to transport (reviewed by Hammes and Schmidt,
2009).
Processes which may influence biochar fate in soil might be the same as
those for other natural organic matter (NOM), although little experimental
evidence on this is still available. If that is the case, a lower clay content and
an increase in soil temperature and water availability will probably enhance
biochar degradation and loss, as previously suggested by Sohi et al. (2009).
For example, mean annual temperature of the site that biochar is applied to
has shown to be a contributing factor in accelerating biochar oxidation in soil
(Cheng et al., 2008). One could hypothesize that the same might apply to
tillage (Sohi et al., 2009) through altering soil aggregate distribution.
Interestingly, Brodowski et al. (2006) did not find evidence that different
management practices have an effect on BC contents in Haplic Luvisol topsoil
(0-30 cm; 13.4±0.2 g kg-1 organic C) from continuous wheat and maize plots.
Adjacent grassland (0-10 cm; 10.3 g Kg-1 organic C; since 1961) and spruce
forest (0-7 cm; 41.0 g kg-1 OC; since ca. 1920) topsoil were also sampled
(Brodowski et al., 2006).
Sohi et al. (2009) and Collision et al. (2009) proposed that feedstock material
(including its degree of aromaticity) and cropping patterns (which influences
nutrient composition in the rhizosphere) are contributing factors in determining
biochar degradation rates in soil. These authors provided the following
example: Pyrolysis of wood-based feedstocks generate coarser and more
resistant biochars explained by the rigid xylemic structure of the parent
material, whereas biochars produced from crop residues (e.g. rye, maize) and
manures are generally finer and nutrient-rich, therefore more readily
degradable by microbial communities (Collison et al., 2009).
Cheng et al. (2008) have recently assessed the effects of climatic factors
(mainly temperature) on biochar oxidation in natural systems. The cation
exchange capacity of biochar was correlated to the mean temperature and the
extent of biochar oxidation was related to its external surface area, being
seven times higher on the external surfaces than in its interior (Cheng et al.,
2008). In addition, X-ray photoelectron spectroscopy (Cheng et al., 2006) and
later, near-edge X-ray absorption fine structure spectroscopy (Lehmann et al.,
2005) have shown that abiotic oxidation occurs mainly in the porous interior of
biochar, while biotic oxidation is the predominant process on external
surfaces. This probably means that biotic oxidation may become more
relevant as particle size decrease as a consequence of biochar weathering,
although there are doubts on the relative importance of such a process
(Cheng et al., 2006). Nevertheless, the influence of increasingly warmer

71
climates on biochar degradation rates in natural systems has not been
resolved yet.
Translocation of biochar within the soil profile and into water systems may
also be a relevant process contributing to explain biochar loss in soil
(Hockaday et al., 2006). Such a translocation via aeolian (e.g. Penner et al.,
1993) and mostly fluvial (e.g. Mannino and Harvey, 2004) long-range
transport has been previously proposed for other forms of BC, in order to
explain its occurrence in deep-sea sediments (Masiello and Druffel, 1998), as
well as in natural riverine (Kim et al., 2004) and estuarine (Mannino and
Harvey, 2004) water.
Soil erosion (in a global context) might result in greater amounts of BC being
redistributed onto neighbouring hill slopes and valley beds (Chaplot et al.,
2005), or enriching marine and river sediments through long-range transport,
as recently suggested by Rumpel et al. (2006a;b) for tropical sloping land
under slash and burn agriculture. Partially, this can be explained by the light
nature (low mass) of biochar (Rumpel et al., 2006a;b), and may be particularly
relevant for finer biochars or those with higher dust contents. Similarly, this
might apply predominantly to soils and sites which are more prone to erosion
(Hammes and Schmidt, 2009).
Up to now, biochar loss and mobility through the soil profile and into the water
resources, has been scarcely quantified and translocation mechanisms are
poorly understood. This is further complicated by the limited amount of long-
term studies and the lack of standardized methods for simulating biochar
aging and for long-term environmental monitoring (Sohi et al., 2009). Sound
knowledge at this level will not only enable for a more robust estimate of
global BC budget to be put forward (through an improved understanding of
the role of BC as a global environmental carbon sink) but also attenuate
uncertainties in relation to current estimates of BC environmental fluxes.
The finest biochar dust fraction, comprising condensed aromatic carbon in the
form of fullerene-like structures (Harris, 1997), is thought to be the most
recalcitrant portion of the BC continuum in natural systems (Buzea et al.,
2006). Interactions between this ultrafine fraction and soil organic and mineral
surfaces has been suggested to contribute to biochar’s inherent recalcitrance
(Lehmann et al., 2009), although quantifying its relative importance by
experimental evidence, may render difficult. Free sub-micron BC particles are
primarily transported to the oceans, where the majority is deposited on coastal
shelves, while smaller amounts continue on to deep-ocean sediments
(Masiello and Druffel, 1998; Mannino and Harvey, 2004) with expected
residence times of thousands of years (Masiello and Druffel, 1998). The
remaining fraction remains suspended in the atmosphere in the form of
aerosols (Preston and Schmidt, 2006) and can be transported over long
distances, eventually reaching the water courses and sediments (Buzea et al.,
2006).

3.2.2 Sorption of Hydrophobic Organic Compounds (HOCs)

The sorption of anthropogenic hydrophobic organic compounds (HOC) (e.g.
PAHs, polychlorinated biphenyl - PCBs, pesticides and herbicides) in soils
and sediments, is generally described based on two coexisting and

72
simultaneous processes: absorption into natural (amorphous) organic matter
(NOM) and adsorption onto occurring charcoal materials (Cornelissen et al.,
2005; Koelmans et al., 2006). Comparatively to that of NOM, charcoals
(including soot) generally hold up to 10-1000 times higher sorption affinities
towards such compounds (Chiou and Kile, 1998; Bucheli and Gustafsson,
2000, 2003). It has been estimated that BC can account for as much as 80-
90% of total uptake of trace HOC in soils and sediments (Cornelissen et al.,
2005), and that it applies to a much broader range of chemical species than
previously thought (Bucheli and Gustafsson, 2003; Cornelissen et al., 2004).
Biochar application is, therefore, expected to improve the overall sorption
capacity of soils (Chiou 1998), and consequently, influence toxicity, transport
and fate of trace contaminants, which may be already present or are to be
added to soils. Enhanced sorption capacity of a silt loam for diuron (Yang and
Sheng, 2003) and other anionic (Hiller et al., 2007) and cationic (Sheng et al.,
2005) herbicides has previously been reported following the incorporation of
biochar ash from crop (wheat and rice) residues. The relative importance of
these latter studies is justified by the fact that charring of crop residues is a
widespread agricultural practice (Hiller et al., 2007). Nevertheless, while the
feasibility for reducing mobility of trace contaminants in soil might be
beneficial (see Section 4.3), it might also result in their localised accumulation,
with potentially detrimental effects on local flora and fauna if at some point in
time the sorbed compounds become available to organisms. Experimental
evidence is required to verify this.
Despite that little is still known on the micro-scale processes controlling
sorption to biochar (Sander and Pignatello, 2005) in soils and sediments, it
has been suggested that it is mechanistically different from the traditional
sorption models for NOM, and that it is also a less reversible process
(Gustafsson et al., 1997; Chiou and Kile, 1998; Jonker et al., 2005). While
absorption to NOM has little or no concentration dependence, adsorption to
biochars has been shown to be strongly concentration dependent (e.g.
Gustafsson et al., 1997; Sander and Pignatello, 2005; Pastor-Villegas et al.,
2006; Wang et al., 2006; Chen et al., 2007), with affinity decreasing for
increasing solute concentrations (Cornelissen et al., 2004; Wang et al., 2006).
Several equations have been employed to describe such a behaviour,
including that of Freundlich (e.g. Cornelissen et al., 2004) and Langmuir (e.g.
van Noort et al., 2004), although more recent equations based on pore-filling
models have shown better fits (e.g. Kleineidam et al., 2002).
Previous studies have convincingly demonstrated that adsorption to charcoals
is mainly influenced by the structural and chemical properties of the
contaminant (i.e. molecular weight, hydrophobicity, planarity) (Cornelissen et
al., 2004, 2005; Zhu and Pignatello, 2005; Zhu et al., 2005; Wang et al.,
2006), as well as pore size distribution, surface area and functionality of the
charcoal (e.g. Wang et al., 2006; Chen et al., 2007). For example, sorption of
tri- and tetra-substituted-benzenes (such as trichlorobenzene, trinitrotoluene
and tetramethilbenzene) to maple wood charcoal (400°C) was sterically
restricted, when comparing to that of the lower size benzene and toluene (Zhu
and Pignatello, 2005). Among most classes of common organic compounds,
biochar has been shown to adsorb PAHs particularly strongly, with desorption
having been regarded as ‘very slow’ (rate constants for desorption in water of

73
10-7-10-1 h-1, and even lower in sediments) (Jonker et al., 2005). This can be
explained both by the planarity of the PAH molecule, allowing unrestricted
access to small pores (Bucheli and Gustafsson, 2003; van Noort et al., 2004),
and the strong π-π interactions between biochar’s surface and the aromatic
molecule (e.g. Sander and Pignatello, 2005). ). In fact, experimental evidence
has recently demonstrated that organic structures in the form of BC (including
biochar) or NOM, which are equipped with strong aromatic π-donor and -
acceptor components, are capable of strongly adsorbing to other aromatic
moieties through specific sorptive forces other than hydrophobic interactions
(Keiluweit and Kleber, 2009).
Although a large body of evidence is available on the way the characteristics
of HOC influence sorption to biochars, the contribution of the char’s properties
to that process has been far less evaluated. It is generally accepted that
mechanisms leading to an increase in surface area and/or hydrophobicity of
the char, reflected in an enhanced sorption affinity and capacity towards trace
contaminants, as demonstrated for other forms of BC (Jonker and Koelmans,
2002; Noort et al., 2004; Tsui and Roy, 2008). The influence of pyrolysis
temperatures mostly in the 340-400°C range (James et al., 2005; Zhu et al.,
2005; Tsui and Roy, 2008) and feedstock type (Pastor-Villegas et al., 2006)
on such a phenomena has been recently evaluated for various wood chars by
a number of authors. Interestingly, sorption to high-temperature chars appear
to be exclusively by surface adsorption, while that to low-temperature chars
derive from both surface adsorption and (at a smaller scale) absorption to
residual organic matter (Chun et al., 2004).
The influence of micropore distribution on sorption to biochars has been
clearly demonstrated by Wang et al. (2006). Diminished O functionality on the
edges of biochar’s graphene sheets due to heat treatment (e.g. further
charring), resulted in enhanced hydrophobicity and affinity for both polar and
apolar compounds, by reducing competitive adsorption by water molecules
(Zhu et al., 2005; Wang et al., 2006). The treated char also revealed a
consistent increase in micropore volume and pore surface area, resulting in
better accessibility of solute molecules and an increase in sorption sites
(Wang et al., 2006).
Once released in the environment, the original adsorption properties of
biochar may be affected by ‘aging’ due to environmental factors, such as the
impact of coexisting substances. The presence of organic compounds with
higher hydrophobicity and/or molecular sizes have shown reduce adsorption
of lower molecular weight compounds to biochars (e.g. Sander and Pignatello,
2005; Wang et al., 2006). In the same way, some metallic ions (e.g. Cu2+,
Ag+) present at environmental relevant concentrations (50 mg L-1) may
significantly alter surface chemistry and/or pore network structure of the char
through complexation (Chen et al., 2007).
Perhaps a more important mechanism to consider, is the influence of
dissolved NOM, including the humic, fulvic (Pignatello et al., 2006) and lipid
(Salloum et al., 2002) fractions, on the physical-chemical properties and
adsorption affinity and capacity of biochars (Kwon and Pignatello, 2005).
Similar evidence has long been reported for activated carbon (Kilduff and
Wigton, 1999). “Aging” of maple wood charcoal (400°C) particles in a

74
suspension of Amherst peat soil (18.9% OC)-water has demonstrated that
NOM reduced affinity of the char for benzene (Kwon and Pignatello, 2005),
corroborating other research (Cornelissen and Gustafsson, 2005; Pignatello
et al., 2006). Similar observation over a period of 100 years has been
reported for pyrene in forest soil enriched with charcoal (Hockaday, 2006). In
both cases, such a behaviour was explained by mechanisms of pore blockage
(Kwon and Pignatello, 2005; Pignatello et al., 2006), and by the capacity of
NOM to compete with (e.g. Cornelissen and Gustafsson, 2005) and displace
the organic compound from the sorption sites (Hockaday, 2006). A wider
range of soil characteristics remain to be tested.
Frequently, contaminated soils contain a mix of organic solvents, PAHs,
heavy metals and pesticides, adding to the naturally occurring mineral and
organic matter (Chen et al., 2007). Nevertheless, most studies on organic
sorption to charred materials have relied on single-solute experiments,
whereas those using multiple solutes hold more practical relevance (Sander
and Pignatello, 2006). Competitive sorption can be a significant environmental
process in enhancing the mobility as well as leaching potential of HOC in
biochar-enriched soil.
Most of the evidence of increased sorption to HOC by biochar incorporation
into soil is indirect (i.e., bulk and biochar or soot sorption is determined
separately and biochar’s contribution is then proved comparatively to a
treatment without biochar) and earlier attempts for its direct assessment
overestimated it (Cornelissen and Gustafsson, 2004). Yet, the potential of
biochar amendment of soils for enhancing soil sorption capacity and,
therefore mitigating the toxicity and transport of relevant environmental
contaminants in soils and sediments appears undeniable. One can suggest
that such an enhancement of soil sorption capacity may result in long mean
residence times and accumulation of organic contaminants with potentially
hazardous health and environmental consequences. At this stage, very little is
known about the short- and long-term distribution, mobility and bioavailability
of such contaminants in biochar-enriched soils.
It is worth underlining that although such a strong adsorptive behaviour
appears to imply a reduced environmental risk of some chemical species (e.g.
PAHs), very little data is, in fact, currently available which confirms this. The
underlying sorption mechanism, including the way it is influenced by a wide
range of factors inherent to the contaminant, to the char material and to the
environment, remains far from being fully understood (Fernandes and Brooks,
2003), and thus it is identified in this report as a priority for research. In this
context, it is vital to comprehensively assess the environmental risk
associated to these species in biochar-enriched soils, while re-evaluating both
the use of generic OC-water distribution coefficients (Jonker et al., 2005) and
of remediation endpoints (Cornelissen et al., 2005). For instance, remediation
endpoints (undetectable, non-toxic or environmentally acceptable
concentrations, as set by regulatory agencies) for common environmental
contaminants in biochar-enriched soils would need to be assessed based on
dissolved (bioavailable) concentrations rather than on total concentrations
(Pointing, 2001; Cornelissen et al., 2005). In order to achieve that, prior
careful experimental evaluation of the contaminant distribution, mobility and
availability in the presence of biochar is paramount.

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3.2.3 Nutrient retention/availability/leaching
Reduction of nutrient leaching from agriculture is an objective in line with the
Water Framework Direct (WFD). The WFD promotes an integrated
management approach to improve the water quality of European water
bodies. Application of fertilisers has led to increased concentrations of nitrates
and phosphates in European surface and ground waters. Specific water
quality targets have been set by the Water Framework Directive with respect
to nitrates, which are very susceptible to leaching (European Parliament and
the Council of the European Union, 2000). Improved agricultural management
practices are increasingly stimulated by the Common Agricultural Policy (cf.
CAP Health Check).
Evidence from several laboratory and field studies suggests that the
application of biochar may lead to decreased nutrient leaching (studies
particularly focussed on nitrates) and contaminant transport below the root
zone. Several mechanisms contribute to the decrease in nutrient leaching
which are related to increased nutrient use efficiency by increased water and
nutrient retention (residence time in the root zone) and availability, related to
an increased internal reactive surface area of the soil-biochar matrix,
decreased water percolation below the root zone related to increased plant
water use (increased evaporative surface), and increased plant nutrient use
through enhanced crop growth. Higher retention times also permit a better
decomposition of organic material and promote the breakdown of
agrichemicals. Nevertheless, mechanisms such as colloid-facilitated transport
of contaminants by biochar particles, or preferential flow induced by biochar
applications, and long term stability of biochar in soil, are potential factors that
my increase the leaching of nutrients and/or contaminants.
The magnitude and dynamics resulting from biochar application are time,
space and process specific. The myriad of interactions within the soil-plant-
atmosphere, and the range of potential feedstock specific effects of biochar
on these interactions, makes it inherently difficult to formulate generic qualities
of “biochar”. It also has to be kept in mind that other factors, such as rainfall
patterns and agricultural management practices, will be more strongly
determining the loss of nutrients from the root zone.
The mobility of the water percolating beyond the root zone depends on the
infiltration capacity, hydraulic conductivity and water retention of the root
zone, the amount of crop transpiration dependent on the density and
capability of the root network to extract water, and the prevalent
meteorological conditions at the site. These factors are largely dependent on
the proportion and connections between micro, meso and macro pores.
The partitioning of groundwater recharge, surface-water runoff and
evapotranspiration is affected by changes in the soil’s water retention
capacity. In those situations where biochar application improves retention (of
plant available water) and increases plant transpiration (Lehmann et al.,
2003), percolation below the root zone can be reduced, leading to the
retention of mobile nutrients susceptible to leaching such as nitrates, or base
cations at low pH.
Biochar directly contributes to nutrient adsorption through charge or covalent
interactions on a high surface area. Major et al. (2002) showed that biochar

76
must be produced at temperatures above 500°C or be activated to results in
increased surface area of the biochar and thus increased direct sorption of
nutrients. Glaser et al. (2002) conclude that ‘charcoal may contribute to an
increase in ion retention of soil and to a decrease in leaching of dissolved OM
and organic nutrients’ as they found higher nutrient retention and nutrient
availability after charcoal additions to tropical soil. A possible contributing
mechanism to increased N retention in soils amended with biochar is the
stimulation of microbial immobilisation of N and increased nitrates recycling
due to higher availability of carbon (see Section 3.2.3). Biological N fixation by
common beans was reported to increase with biochar additions of 50 g kg-1
soil (Rondon et al., 2007), although soil N uptake decreased by 50%, whereas
the C:N ratios increased with a factor of two.
Lehmann et al. (2003) reported on lysimeter experiments which indicated that
the ratio of uptake to leaching for all nutrients increases with charcoal
application to the soil. However they also concluded that it could not clearly be
demonstrated which role charcoal played in the increased retention, although,
in these experiments, water percolation was not decreased. Therefore,
nutrients must have been retained on electrostatic adsorption complexes
created by the charcoal. Similarly, Steiner et al. (2004) attributed decreased
leaching rates of applied mineral fertiliser N in soils amended with charcoal to
increased nutrient use efficiency. Nevertheless, the interaction between
mineral fertiliser and biochar seems critical. Lehmann et al. (2003) found that
while cumulative leaching of mineral N, K, Ca and Mg in an Amazonian Dark
Earth was lower compared to a Ferralsol in unfertilised experiments, leaching
from the ADE exceed that from the Ferralsol in fertiliser experiments.
If biochar applications lead to improved soil aggregation, this may lead to an
increase in the soil’s water infiltration capacity. Using measured properties
such as saturated hydraulic conductivity and total porosity in a modelling
assessment of the impact of charcoal production, Ayodele et al. (2009)
showed that infiltration was enhanced and runoff volume reduced. The
increase in infiltration may be accompanied by improved water retention in the
root zone in coarse soils. On the other hand, however, since a large
percentage of the pores in biochar are very small (<2 x 10-3 μm, following
Tseng and Tseng, 2006), it may also reduce the mobility of water through the
soil. If the increased infiltration is not off-set by increased retention and
transpiration, due to factors related to the native soil, and/or if crop nutrient
uptake is not increased, the net results may be an increased percolation
below the root zone, especially of soluble and mobile nutrients such as
nitrates.
Fine biochar particles resulting from transportation, application, and further
weathering in the field, may facilitate the colloidal transport of nutrients and
contaminants (Major et al., 2002).
Hydrophobicity (see Section 3.2.2) induced by biochar is thought to be most
significant in the first years after application since ‘fresh’ biochar contains a
large fraction of hydrophobic groups. The implications of biochar
hydrophobicity on runoff and unwanted export of nutrients from the field has
not been investigated in detail. Another potential concern in certain soils is
preferential flow induced by the incorporation of biochar in the soil matrix, it

77
has been suggested that biochar can alter percolation patterns, residence
times of soil solution, and affect flow paths (Major et al., 2002).

3.2.4 Contamination
Given that the widespread interest in biochar applications to soils continues to
rise, so does the concern regarding the potential for soil contamination
associated to some of its components. It is crucial to ensure that soil functions
and processes as well as water quality are not put at risk as a consequence of
biochar application to soils, which would carry severe health, environmental
and socio-economic implications (Collison et al., 2009). Mineral contaminants
like salts that are often present in some biochars and may be detrimental to
soil functioning rather than to human and animal health, and have been
discussed previously. This section is dedicated to contaminants such as
heavy metals, PAHs and dioxins, which remain major issues of concern with
regard to potential for soil contamination and health hazards, and yet have
surprisingly received very little attention.
The occurrence of these compounds in biochar may derive either from
contaminated feedstocks or from pyrolysis conditions which favour their
production. For example, slow pyrolysis at temperatures below 500°C is
known to favour the accumulation of readily available micronutrients (e.g
Sulphur) in biochar (Hossain et al., 2007). However, heavy metals, PAHs and
other species with disinfectant and antibiotic properties (e.g. formaldehydes,
creosols, xylenols, acroleyn) may also accumulate under such operating
conditions (Painter, 2001). Full and careful risk assessment for such
contaminants is urgently required, in order to relate contaminant toxicity to
biochar type, safe application rates and operating pyrolysis conditions.
Organic wastes (e.g. biosolids, sewage sludge, tannery wastes) are known to
generally contain high levels of light and heavy metals, which remain in the
final biochar product following pyrolysis (Hospido et al., 2005; Chan and Xu,
2009). Bridle and Pritchard (2004) reported high concentrations of Copper
(Cu), zinc (Zn), chromium (Cr) and nickel (Ni) in biochar produced from
sewage sludge. Muralidhar (1982) has long found that Cr, which accounts for
up to 2% (total dry weight) of tannery wastes, is commonly found in biochar
produced from this material. On the other hand, relatively low concentrations
of aluminium (Al), Cr, Ni and molybdenum (Mo) have been recently detected
in poultry litter, peanut hull and pine chip biochars produced between 400-
500°C, while poultry litter biochar generally contained the highest levels of
these metals (Gaskin et al., 2008). In contrast, Zn, Cu, Al and Fe were lower
in the poultry litter biochar compared to that in pine chip and peanut hulls
biochars, which pattern seem to be reverse to that observed in the feedstock
materials. Although one could suggest pyrolysis as means of reducing metal
availability in some feedstocks (such as poultry litter), and be encouraged to
use biochar (instead of poultry litter) for mitigating some of the concerns
relating to soil contamination, there is no clear evidence to confirm this
(Gaskin et al., 2008).
Metal concentration in the biomass feedstock often determines biochar’s safe
application rate (McHenry, 2009). Preliminary data seems to suggest that, at
current ordinary biochar application rates, there is little environmental risk by
metal species within biochar, which McHenry (2009) describes as similar to

78
that associated to the use of conventional fertilisers. In fact, for contaminants
such as Zn, mercury (Hg), arsenic (Ar), lead (Pb) and Ni, it is likely that
significant risk can only be expected from exceedingly high biochar
application rates (>250 t ha-1) (McHenry, 2009). A wider range of biochars and
soil types remains to be tested, which would undoubtedly shed more light onto
the potential for soil and water contamination by metals.
Secondary chemical reactions during thermal degradation of organic material
at temperatures exceeding 700°C, is generally associated to the generation of
heavily condensed and highly carcinogenic and mutagenic PAHs (Ledesma et
al., 2002; Garcia-Perez, 2008). Nevertheless, little evidence exists that PAHs
can also be formed within the temperature range of pyrolysis (350-600°C),
although these appear to carry lower toxicological and environmental
implications (Garcia-Perez, 2008). Nevertheless, their potential occurrence in
the soil and water environments via biochar may constitute a serious public
health issue. Evidence seems to show that biomass feedstock and operation
conditions are influencing factors determining the amount and type of PAHs
generated (Pakdel and Roy, 1991), and therefore, there is great need to
assess the mechanisms, as well as identify specific operational and feedstock
conditions, which lead to their formation and retention in the final biochar
product.
Very little data is available on the occurrence of PAHs in pyrolysis products,
compared to that from combustion or incineration. Among such studies,
Fernandes and Brooks (2003), Brown et al. (2006) and Jones (2008) do stand
out. Pea straw and eucalyptus wood charcoal produced at 450°C for 1 h,
exhibited low PAHs concentrations (<0.2 µg g-1), although their levels in straw
(0. 12 µg g-1) were slightly higher than that from the denser feedstock material
(0.07 µg g-1) (Fernandes et al., 2003). Similarly, Brown et al. (2006) reported
that PAHs concentrations in several chars produced at temperatures
exceeding 500°C, ranged between 3-16 µg g-1 (depending on peak treatment
temperature), compared to that (28 µg g-1) in char from prescribed burn in
pine forest. The range of producing conditions and feedstock materials
employed in the latter studies was narrow. In contrast, Jones (2008) studied
twelve biochar products from a variety of biomass sources and producers,
with evidence that PAHs levels in biochar were often comparable or even
lower than those found in some rural urban and urban soils. This finding
corroborates previous studies (reviewed by Wilcke, 2000), in which topsoil
concentration ranges of several PAHs were found to increase in the order of
arable < grassland < forest < urban. For example, at the lower end (arable
soil), concentration ranges for naphthalene, fluorene, phenanthrene,
anthracene and pyrene were up to 0.02, 0.05, 0.067, 0.134 µg g-1
(respectively). At the top end of the concentration range (urban soil), levels of
those compounds (respectively) were up to 0.269, 0.55, 2.809, 1.40 and
11.90 µg g-1 (reviewed by Wilcke, 2000). It is important to note, however, that
the latter data refers to initial concentrations in soil, not taking into account
interactions with organic and mineral fractions, and most importantly, not
providing information on the bio-available fraction.
Recently, however, the mild (supercritical fluid) extraction of pyrogenic PAHs
from charcoal, coal and different types of soot, including coal soot, showed
promising results (Jonker et al., 2005). To the best of our knowledge, this

79
study was pioneer in reporting desorption kinetics of pyrogenic PAHs from
their ‘natural’ carrier under conditions which mimic those in natural
environments. Such “soot and charcoal-associated PAHs” were found to be
strongly sorbed to their carrier matrix (e.g. charcoal, soot) by means of
physical entrapment within the matrix nanopores (so called “occlusion sites”)
in charcoal and sequestration within the particulate matter. Consequently, it is
anticipated “very slow desorption” (rate constants of up to 10-7 to 10-6 h-1) of
these compounds from the carrier in natural environments, which can range
from several decades to several millennia (Jonker et al., 2005). PAHs sorption
to charcoals has been reviewed extensively in Section 3.2.2 of this report,
including the mechanisms leading to increases in their accessibility, such as
interactions with NOM and coexisting chemical species.
To the best of our knowledge, there are no toxicological reports involving
PAHs incorporated in soil due to biochar application, nor have biochar
application rates have been defined in terms of PAHs accumulation and
bioavailability, both in soil and water systems. Further research is paramount
on the behaviour of such contaminants in biochar-enriched natural systems.
In this context, a re-evaluation of risk assessment procedures for these
compounds needs to be put in place, which takes into account the influence of
NOM on their desorption from biochar, transport and bioavailability.
Dioxins and furans are planar chlorinated aromatic compounds, which are
predominantly formed at temperatures exceeding 1000°C (Garcia-Perez,
2008). Although data exists confirming their presence in products from
combustion reactions, such as incineration of landfill and municipal solid
wastes (as cited by Garcia-Perez, 2008), no reports were found on their
content in biochar derived from traditional biomass feedstocks. In contrast,
char from automobile shredder residues was shown to contain up to 0.542 mg
kg-1 of dioxins, while their generation and accumulation in the char was
dependent on the operational conditions (Joung et al., 2007). Scarce
experimental evidence on dioxin levels in pyrolysis products (biochar in
particular) in the range of temperatures between 350-600°C, is largely limiting
towards our knowledge on potential dioxin contamination of soil via biochar.
More research on this matter is urgently needed. It appears that pyrolysis of
strongly oxygenated feedstocks under low temperatures (400 and 600°C) do
not favour the generation of dioxins and dioxin-related compounds. Based on
the current knowledge, it is likely that such a risk is low for the aforementioned
biochar production factors, particularly when using low-chlorine and low-metal
containing feedstocks (Garcia-Perez, 2008).
At this stage, extrapolating a link between the presence of contaminants on
biochar and a detrimental effect on human and animal health, particularly in
regard to bioaccumulation and bioamplification in the food chain, can only be
hypothesised. One can suggest that potential uptake and toxicity of such
contaminants is perhaps more prominent in the case of microbial
communities, sediment-dwelling organisms and filter feeders. In note of the
application of biochar into soil being an irreversible process, Blackwell et al.
(2009) emphasised the need for full case-by-case characterisation and risk
assessment of each biochar product previous to its application to soil,
accounting not only for heterogeneity among biochars, but also for soil type
and environmental conditions. There are no current standards for biochar or

80
processing conditions which can provide sound basis for biochar quality
regulations with regard to the presence of contaminants, thus ensuring soil
and water protection. Also lacking is a clearly defined set of conditions under
which biochar and related materials can be applied to soil without licensing
(Sohi et al., 2009).
As Collison et al. (2009) noted, the natural occurrence of BC in soils is
widespread and detrimental effects on environmental quality are generally not
apparent. However, it is the perspective of an extensive and indiscriminate
incorporation of biochars into soils, derived from some feedstock materials
under specific operation conditions, without previous full risk assessment,
which constitutes the main issue of concern. This is particularly the case for
small-scale and on-farm pyrolysis units using local biomass resources (e.g.
forestry and agricultural wastes), which may not hold the necessary
technological and economic infrastructures to tackle this matter. Also, it is
likely that these small landholders in rural areas might prefer using low-
temperature pyrolysis, thereby reducing operation costs. Farmers should be
made aware that sub-optimal pyrolysis operating conditions and certain
feedstocks may not only reduce the benefits associated to biochar application,
but also enhance the risk of land and water contamination.

3.2.5 Soil Organic Matter (SOM) Dynamics

SOM stabilisation mechanisms for temperate soils have been researched
comprehensively and reviewed recently (Von Lützow et al, 2006; 2008 2008;
Kögel-Knabner et al., 2008; Marschner et al., 2008).
Primary recalcitrance refers to the recalcitrance of the original plant matter,
while secondary recalcitrance refers to that of its charred product, i.e.
pyrogenic BC. For biochars from feedstocks that have already undergone
selective preservation, i.e. any process leading to the relative accumulation of
recalcitrant molecules, it may be appropriate to consider tertiary recalcitrance.
Stability of SOM is the result of recalcitrance, organo-mineral interactions, and
accessibility. Because biochar is OM but also has many properties
functionally similar to mineral matter, it is necessary to consider the stability of
biochar in soils as well as the stability of native SOM, or OM that is added
with, or after, the biochar.

3.2.5.1 Recalcitrance of biochar in soils

Studies of charcoal produced by wildfires have shown that abiotic processes
generally have more impact on the decomposition of charcoal than biotic
processes, in the short term (Cheng et al 2006; Bruun and Luxhøi. 2008).
However, abiotic oxidation can only occur on the surface and as such once
the surface of biochar has been oxidised biotic process are thought to
become more important. The fact that the soil microbiota is capable of
oxidising graphitic carbon, which is thermodynamically stable and recalcitrant
carbon, was first demonstrated by Shneour (1966). This author found that a
‘substantially higher’ oxidation rate, being at least a 3-fold increase, was found
in non-sterile soils than in sterilised soils.
More work regarding recalcitrance has been conducted on BC, specifically
pyrogenic BC, rather than on biochar per se. Nevertheless, owing to its
relatively similar physical and chemical composition BC is an acceptable

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analogue and it is likely that the recalcitrance of biochars will function
according to similar mechanisms.
As graphite has been shown to be oxidised by microbial activity, albeit very
slowly (Shneour 1966), a degree of decomposition of biochars can be
expected. Contradictory experimental results exist, with both rapid (Bird et al.
1999) and slow (Shindo 1991) decomposition of biomass-derived BC being
reported. This difference is likely to be an artefact of the different microbial
communities to which the BC was exposed. Although precise details
regarding the turnover of BC in soils remain unknown, and due to the
complexity of its interaction within the soil system and its biota exact details
are unlikely to be found, BC has been found to be the oldest fraction of C in
soil, being older than the most protected C in soil aggregates and organo-
mineral complexes (Pessenda et al., 2001), which are commonly the most
stable forms of C in soil. This demonstrates that even without knowing the
precise details of turnover of BC in soil, it at least has highly stable
components with “decomposition leading to subtle, and possibly important,
changes in the bio-chemical form of the material rather than to significant
mass loss” (Lehmann et al 2006).
It has been noted that the recalcitrance of BC in soils cannot be characterised
by a single number (Hedges et al., 2000; Von Lützow et al., 2006). This is
because pyrogenic BC is an amalgamation of heterogeneous compounds
and, as such, different fractions of it will decompose at different rates under
different conditions (Hedges et al., 2000). According to Preston & Schmidt
(2006) the more recalcitrant compounds in pyrogenic BC, created by wildfire
and therefore of a woody feedstock, can be expected to have a half life in the
region of thousands of years (possibly between 5 and 7 thousand years) in
cold and wet environments. However, some fractions of pyrogenic BC which
may have undergone less thermal alteration (being more analogous to
biochars which have also undergone less thermal alteration due to low heat
pyrolysis, a half life in the region of hundreds of years as opposed to
thousands may be expected (Bird et al., 1999). This agrees with work
reported by Brunn et al. (2008) who found that the rate of microbial
mineralisation of charcoal decreases with increasing mineralisation
temperature (see also Section 1.6).
Besides physical and chemical stabilization mechanisms, another important
factor that may affect the residence time of biochar in soils is the phenomenon
of co-metabolism. This is where biochar decomposition is increased due to
microbial metabolism of other substrates, which is often increased when SOM
is ‘unlocked’ from the soil structure due to disturbance (e.g. incorporating
biochar into the soil via tillage).

3.2.5.2 Organomineral interactions

The interactions between SOM and soil minerals have received considerable
attention in the literature. Von Lützow et al. (2006) concluded that some
evidence exists for interactions between biochar and soil minerals, leading to
accumulation in soil, but that the mechanisms responsible are still unknown.
One potential mechanism is the oxidation of the functional groups at the
surface of the charcoal, which favours interactions with soil organic and

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mineral fractions (Lehmann et al., 2005; Glaser et al., 2002). Section 3.2.1
explores further the interaction between biochar and other soil components.

3.2.5.3 Accessibility
Biochar can both increase and decrease the accessibility of SOM to
microorganisms and enzymes. Brodowski et al. (2006) provided evidence that
a significant portion of BC occurs in the aggregate-occluded OM in soil.
Interestingly, the largest BC concentrations occurred in microaggregates
(<250 µm) and it has been suggested that it may be actively involved in the
formation and stabilisation of microaggregates, comparatively to other forms
of organic matter (Brodowski et al., 2006). At the present, one can only
speculate on such a role of biochar in soil. Most importantly, organo-mineral
interactions may be relevant in determining the environmental behaviour and
fate of biochar (Hammes and Schmidt, 2009; Section 3.2.1) and can
contribute to physically protecting it from degradation, while promoting its long
mean-residence times in soil (Glaser et al., 2002; Lehmann et al., 2005;
Brodowski et al., 2006).

3.2.5.4 Priming effect

The priming effect has been defined as being “the acceleration of soil C
decomposition by fresh C input to soil” (Fontaine et al., 2004) and are
generally considered to be short-term changes in the turnover of SOM
(Kuzyakov et al., 2000). The priming effect is thought to be a function of
changes in microbial community composition upon fresh C input into soil (e.g.
cellulose, Fontaine et al., 2004). This means that addition of a ‘new’ source of
carbon into the soil system can potentially lead to a priming effect whereby
SOC is reduced. Several mechanisms may be involved: changes in pH,
changes in water-filled pore space, changes in habitat structure, or changes in
nutrient availability.
Following cellulose addition, Fontaine et al. (2004) found that decomposition
rate of soil humus stock in savannah soil increased by 55%. Kuzyakov et al.
(2009) demonstrated that BC in soil underwent increased decomposition upon
the addition of glucose to soil. They concluded that while soil microorganisms
were not dependant on BC as an energy source, the extracellular enzymes
produced by the microbial community for the decomposition of the glucose
(and its metabolites) also decomposed the BC, albeit at a vastly decreased
rate when compared to the added glucose. They estimated the mean
decomposition time of black carbon to be in the range of 0.5% per year and
concluded that the mean residence time of back carbon in soil is likely to be in
the range of about 2000 years. This provides some further evidence of
priming effects occurring with regard to mineralisation of C in soils, in this
case BC, upon addition of a substance, in this case glucose. As to whether
the addition of biochar to soil can lead to a priming effect leading to
accelerated mineralisation of SOM is still a matter of debate.
This then leads to the question as to whether biochar addition to soils can
cause a priming effect. Kuzyakov et al. (2000) stated that the most important
mechanisms concerning priming effects are due to increased activity or
quantity of the microbial biomass. Biochar has been shown to increase both of
these factors (Section 3.2.6.1), and as such there is the clear potential for

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biochar to cause a priming effect on SOM. There is a paucity of data on the
possible priming effect of biochar on SOM, but some initial data is available.
Steinbeiss (2009) found that the addition of homogeneous biochars, made
from glucose and yeast to produce N-free biochar and biochar with a N
content of ~5%, respectively. When these biochars were mixed with arable
soils and forest soils in controlled microcosm experiments a clear priming
effect could be observed with between 8% and 12% of carbon from the SOC
pool being lost in 4 months after addition of either type of biochar to either
type of soil. The addition of nitrogen containing biochar to forest soil had the
largest effect (13% loss) with addition of the nitrogen free biochar to arable
soil having the smallest effect (8%). That said, it is important to note that the
controls of both the arable soil and the forest soil which had no biochar
addition but were subject to the same disturbance (sieved to 2 mm and
mixed) also showed a loss of carbon from the SOC pool of 4% and 6%
respectively. This demonstrates that disturbance to the soil which is sufficient
to break up soil aggregates and expose previously protected soil organic
matter to microbial decomposition and mineralisation itself has a strong
priming effect on SOC.
Biochars made from these specific feedstocks are unlikely to be used in
reality particularly as they were almost certainly lacking in micronutrients such
as P and K which would be introduced into the soil with most biochar types.
Also, they were produced by hydrothermal pyrolysis, which is not the most
commonly used or posited method of pyrolysis. This, combined with large
amount of variance seen within each treatment group means that the results
must be extrapolated with caution. However, it appears to be preliminary
evidence that biochars can instigate, or at least increase the priming effect
and accelerate the decomposition of SOC. There is some evidence that the
availability of N in a soil is the main factor affecting the priming effect, with
more available N leading to a reduced priming effect (Neff et al., 2002;
Fontaine, 2007). This suggests that the priming effect could perhaps be
reduced or eliminated though the co-addition of N fertiliser along with biochar.
If biochar components are highly recalcitrant in soil, as evidence suggests,
and its addition, in some scenarios at least, speeds up the decomposition,
and thereby depletion of SOC, soil fertility and the ecosystem services which
it provides may be negatively affected. It is conceivable that, through biochar
addition to soil, it may be possible to increase the level of C in soils beyond
what is found in most given soils on average at the moment. However, if this
is C in the form of a highly recalcitrant substance that does not take part in the
cycling of C in the soil (i.e. biochar) and not the highly chemically complex and
dynamic substance (i.e. humus) and other SOM fractionations, then
ecosystem functioning of soils may well be compromised. This is because it is
well recognised that it is not the presence of C within the soil which is
important for functioning, but rather it is the decomposition of SOC which
drives the soil biota and leads to the provision of ecosystem services. This
was recognised even before Russell (1926) who stated that SOM must be
decomposed before it has ‘served its proper purpose in the soil’. This is
clearly an area that warrents further research.

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3.2.5.5 Residue Removal
One of the often proposed methods of obtaining biomass for use as a
feedstock to make biochar is the removal of crop residues for pyrolysis. The
removal of residues, and the possible associated impacts has already been
discussed extensively from the point of view of biofuels (Wilhelm et al., 2004;
Lal, 2007; Blanco-Canqui and Lal, 2008; Lal, 2009). Removal of crop residues
is associated with increased risk of soil loss by both water and wind erosion
with associated off-site effects, depletion of SOM, degradation of soil quality
leading to decrease in agronomic productivity and a reduction in crop yields
per unit input of fertiliserand water, thereby compromising the sustainability of
agriculture (Lal, 2007).
Removal of crop residues for biochar production, therefore, has the potential
to have multiple negative effects on the soil, which may only be partially
outweighed, if at all, by the positive effects of biochar addition. While it is
possible that the inclusion of biochar into the soil system may aid the
reduction of atmospheric CO2, it is also feasible that more CO2 will be
required to be produced as a by-product of processes undertaken to
remediate the damage done by crop residue removal, such as increased
production of fertiliser which may need to be undertaken to keep yields stable.
Furthermore, as discussed above, the soil biota relies on the breakdown of
SOM to provide energy for it to perform the multitude ecosystem services
which it provides. It is the SOM dynamics that helps drive the system, not just
the presence of SOM. If the potential new inputs of SOM, being crop residues
in many agricultural situations, are removed, and converted into a
substantially more recalcitrant form which does not function as an energy
source for the edaphic microflora and fauna, then ecosystem services may
well be compromised and reduced.

3.2.6 Soil Biology

The soil biota is vital to the functioning of soils and provides many essential
ecosystem services. Understanding the interactions between biochar, when it
is used as a soil amendment, and the soil biota is therefore vital. It is largely
through interactions with the soil biota, such as promoting arbuscular
mycchorizal fungi (AMF) as well as influences on water holding capacity,
which leads to the reported effects of biochar on yields (see Section 3.3).
Soil is a highly complex and dynamic habitat for organisms, containing many
different niches due to its incredibly high levels of heterogeneity at all scales.
On the microscale, soil is often an aquatic habitat, as micropores in soil are
full of water at all times, apart from very extreme drought, due to the high
water tension which exists there. This is vital for the survival of many microbial
species which require the presence of water for mobility as well as to function.
Indeed, many soil organisms, specifically nematodes and microorganisms
such as protozoa enter a state of cryptobiosis, whereby they enter a
protective cyst form and all metabolism stops in the absence of water. When
biochar application leads to an increased water retention of soils (see Section
3.1.2), it seems likely, therefore, that this will have a positive effect on soil
organism activity, which may well lead to concurrent increases in soil
functioning and the ecosystem services which it provides.

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Organisms in the soil form complex communities and food webs and engage
in many different techniques for survival and to avoid becoming prey, ranging
from hiding in safe refuges, through to conducting forms of chemical ‘warfare’.
Biochar, due to its highly porous nature, has been shown to provide increased
levels of refugia where smaller organisms can live in small spaces which
larger organisms cannot enter to prey on them. Microorganisms within these
micropores are likely to be restriced in growth rate due to relying on diffusion
to bring necessary nutrients and gases, but as this occurs in micropores
within the soil, this demonstrates that microorganisms utilising these refugia
almost certainly would not be reliant of decomposition of the biochar for an
energy source. This is likely to be one of the mechanisms for the
demonstrated increases in microbial biomass (Steiner et al., 2008; Kolb et al.,
2009), and combined with the increased water holding potentials of soil is a
possible mechanisms for the increased observed basal microbial activity
(Steiner et al., 2008; Kolb et al., 2009). However, due to the complexities of
the soil system and its biota, it is probable that many more mechanisms are at
work. For example Kolb et al. (2009) demonstrated that while charcoal
additions affected microbial biomass and microbial activity, as well as nutrient
availability, differences in the magnitude of the microbial response was
dependent on the differences in base nutrient availability in the soils studied.
However, they noted that the influences of biochar on the soil microbiota
acted in a relatively similar way in the soils they studied, albeit at different
levels of magnitude, and so suggested that there is considerable predictability
in the response of the soil biota to biochar application.
As with all interactions between the soil biota and biochar, there is a scarcity
of data regarding the interaction of biochar with fungi. However, considering
the diverse saprophytic abilities of fungi it is probable that the interaction
between fungi and biochar is most likely to affect the stability and longevity of
biochar within the soil. While there is evidence of long residence times of
biochar in soils from Terra Pretas, biochar from different sources and exposed
to different fungal communities may well have differing levels of recalcitrance
and hence residence times in soils. This is therefore a highly pertinent area
for further research.
There is some evidence that the positive effects of biochar on plant production
may be attributable to increased mycorrhizal associations (Nisho and Okano,
1991). The majority of studies concerning biochar effects on mycorrhiza show
that there is a strong positive effect on mycorrhizal abundance associated
with biochar in soil (Harvey et al., 1976; Ishii and Kadoya, 1994; Vaario et al.,
1999). The possible mechanisms were hypothesised by Warnock et al. (2007)
to include (in decreasing order of currently available experimental evidence)
a) alteration of soil physico-chemical properties
b) indirect effects on mycorrhizae through effects on other soil
microbes
c) plant–fungus signalling interference and detoxification of
allelochemicals on biochar
d) provision of refugia from fungal grazers

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Biochar, immediately after pyrolysis, can have a wide range of compounds on
its surface. These can include ones that are easily metabolised by microbes,
such as sugars and aldehydes which are turned over quickly, but may also
include compounds which have bactericidal and fungicidal properties such as
formaldehyde and cresols (Painter, 2001). However, residence times of these
substrates has been shown to be in the range of one to two seasons and,
therefore, long term effects of these chemicals on the soil biota are unlikely
(Zackrisson et al., 1996).
The structure of biochar provides a refuge for small beneficial soil organisms,
such as symbiotic mycorrhyzal fungi which can penetrate deeply into the pore
space of biochar and extraradical fungal hyphae (fungal hyphae which are
found outside of roots) which sporulate in the micropores of biochar where
there is lower competition from saprophytes (Saito and Marumoto, 2002).
Nishio (1996) stated that "the idea that the application of charcoal stimulates
indigenous arbuscular mycorrhiza fungi in soil and thus promotes plant growth
is relatively well-known in Japan, although the actual application of charcoal is
limited due to its high cost". The specifics of the cost-benefit relationship of
biochar application to soil and its associated effects on yield have not yet
been covered in depth by the scientific community and is subject of discussion
in Section 5.4.
The relationship between mycorrhizal fungi and biochar may be important in
realising the potential of charcoal to improve fertility. Nishio (1996) also
reported that charcoal was found to be ineffective at stimulating alfalfa growth
when added to sterilised soil, but that alfalfa growth was increased by a factor
of approximately 1.8 when unsterilised soil containing native mycorrizal fungi
was also added. This demonstrates that it is the interaction between the
biochar and the soil biota which leads to positive effects on yield, and not just
the biochar itself (See Section 3.3).

3.2.6.1 Soil microbiota

It has long been assumed that soil biodiversity and SOM are positively
correlated although experimental evidence for this is scarce. Even if this
assumption is proven to be true, it is unclear as to what role biochar will play
in this interaction. This is because, for the majority of the soil biota at least,
biochar appears to function more as the mineral constituent of the soil, than
the OM per se. Nevertheless, there is experimental evidence that microbial
communities are directly affected by the addition of biochar to soils (Ogawa,
1994; Rondon et al., 2007; Warnock et al., 2007; Steiner et al., 2008).
Due to the fact that experiments involving the addition of biochar to soils are
relatively new, with only relatively few experiments being more than a decade
old, quantifying the long term effects of biochar addition to soil is problematic.
While not perfect analogies to the addition of biochar to temperate soils,
investigation of Terra Preta soils in the Amazon Basin does have the potential
to lead to insights regarding the long term effects of biochar addition to soil.
O’Neill et al. (2009) performed 16s rRNA analysis on Terra Pretas and their
surrounding soils. Although their experiment was limited by the fact that they
isolated microorganisms through culturing techniques, they did find numerous
differences between Terra Pretas and their surrounding soils. Firstly, higher

87
numbers of culturable bacteria, by over two orders of magnitude were found in
the Terra Pretas consisting of five possible new bacterial families. They also
reported greater diversity being isolated from the Terra Preta soils. This
increase in culturable bacterial populations and a greater culturable diversity
were found in all of the Anthrosols, to a depth of up to 1 m, when compared to
adjacent soils located within 50-500 m of terra preta. Although using culturing
of the microorganisms as a form of isolation is undoubtedly a weakness in this
experiment design owing to the fact that the vast majority of soil
microorganisms are not culturable in the laboratory (Torsvik et al., 1990; Ritz,
2007), soil extract media was used which when compared to standard culture
media revealed an increased diversity in the soil microbial populations of the
Terra Pretas
As well as affecting the inherent recalcitrance of biochar, the pyrolysis
temperature range also affects how the biochar will interact with the soil
community. This is particularly true of woody charcoal which, at lower
pyrolysis temperatures retains an interior layer of bio-oil which is equal to
glucose in its effect on microbial growth (Steiner, 2004). When pyrolysed at
higher temperatures, this internal layer of bio-oil is lost and so it is likely that
the biochar will have less impact with regard to promoting soil fertility when
compared to biochar which does have the internal layer of bio-oil.
When added to soil, biochar has been shown to cause a significant increase
in microbial efficiency as a measure of units of CO2 released per microbial
biomass carbon in the soil as well as a significant increase in basal respiration
(Steiner et al., 2008). Steiner et al. (2008) also found that the addition of
organic fertiliser amendments along with biochar lead to further increases in
microbial biomass, efficiency in terms of CO2 release per unit microbial
carbon, as well as population growth and concluded that biochar can function
as valuable component of the soil system, especially in fertilised agricultural
systems.
As well as increasing basal respiration and microbial efficiency, there is
experimental evidence that biochar addition to soil increases N2 fixation by
both free living and symbiotic diazotrophs (Ogawa, 1994; Rondon et al.,
2007). Rondon et al. (2007) reported that the positive effects of biochar,
including increased N2 fixation, lead to a between 30 and 40% increase in
bean (Phaseolus vulgaris L.) yield at biochar additions of upto 50 g kg-1.
However, they found that at an application rate of 90 g kg-1 a negative effect
with regard to yield occurred. It should be noted that this appears contrary to
data shown in Figure 3.1 which shows a general trend for positive crop
productivity effects upon biochar addition to soil. This may be due to the
Rondon et al. (2007) study being excluded from the meta-analysis owing to
the study not reporting the variance of within their treatments meaning that the
data could not be included. This means that a possible negative weighting
was not included in the meta-analysis which could have caused a slight scew
of the results. However, as n was low in the Rondon study when compared to
the combined data used in the meta-analysis, the effects of this omission are
likely to be minimal and this highlights the need of accurate reporting of
variances in experimental data to both allow effective interpretation of the
results, and to allow further analyses such as statistical meta-analyses to be
undertaken. Furthermore, many more studies which are reported in the meta-

88
analysis showed a positive effect on crop productivity at similar or higher
application rates. However, this highlights the fact that while biochar addition
to soil is potentially positive with regard to crop yield, situations also exist
where negative effects can occur regarding yield. There is currently no clear
mechanism which may lead to positive effects on yield can become negative
once a threshold has been crossed regarding the amount of biochar which is
added to soil. While it is possible to hypothesise mechanisms responsible for
this effect, there is, as yet, no experimental evidence to confirm or refute any
hypothesis and this highlights the need for further research.

3.2.6.2 Soil meso and macrofauna

There is a current paucity of research with regard to the interaction of biochar
with the soil meso and macrofauna, with the exception of earthworms.
Both the application rate of biochar and the original feedstock used have been
shown to affect the soil biota. Weyers et al. (2009) reported that application
rates higher than 67 t ha-1 of biochar made from poultry litter had a negative
impact on earthworm survival rates. They hypothesised that increased soil pH
or salt levels may have been the reason for the observed reduced survival
rates. They noted that earthworm activity was greater in soil amended with
pine chip biochar than with poultry litter biochar and so concluded that
different types of biochar can have different effects on the soil biota. This
confirmed work reported by Chan et al. (2008) who found that earthworms
had different preferences for different types of biochar, but noted that the
underlying mechanisms driving these preferences required further work.
Recent work by Van Zwieten et al. (2009) has shown that earthworms
preferred biochar-amended Ferrosols over control soils, although they found
no significant difference for Calcarosols. This shows that it is not just the
application rate or feedstock of the biochar which is important to consider
when predicting possible effects, but the soil to which it is added must also be
taken into account. This highlights the complex dynamic interactions which
can vary greatly with soil type, application rate and feedstock used and shows
that predicting the effects of biochar application on the soil biota of a given
soil, whilst very important, is inherently very difficult.
Some work has been undertaken looking at the effects of charcoal ingestion
on earthworms (Hayes, 1983). When charcoal is ingested by an earthworm,
along with other soil particles, the two are mixed with mucus secreted in the
oesophagus and finely ground in the muscular gizzard. When excreted, the
charcoal/soil paste is stabilised by Van der Waals forces after drying and
forms a dark-coloured humus (Hayes, 1983). Ponge et al. (2006) reported that
in laboratory experiments the earthworm Pontoscolex corethrurus was found
to prefer to ingest a mixture of charcoal and soil compared to either pure soil
or pure charcoal. Because of this, Ponge et al (2006) concluded that
Pontoscolex corethrurus was the organisms most responsible for the
incorporation of charcoal into the topsoil in the form of silt size particles which
aids the formation of stable humus in Terra Pretas.
In further laboratory experiments on the effects of charcoal on populations of
earthworms, Topoliantz and Ponge (2003) found differences in the way in
which different populations of the earthworm species P. corethrurus, taken

89
from either forest soil or fallow soil, were adapted to the presence of charcoal,
implying that the addition of charcoal to soil is exerting a selective influence
on the worms although what the specific effects of this selective pressure may
eventually be is unclear. They also reported that the observed transport of
charcoal within the soil demonstrated the importance of P. corethrurus in the
incorporation of charcoal particles into the soil.
No research has yet been undertaken investigating the effects of biochar
addition to soil on soil microarthropods such as collembola or acari, or on
other soil dwelling organisms such as rotifers and tardigrades. Any negative
effect on these organisms seems likely to only occur as a result of any
contamination which exists in the biochar, if that contaminant is bioavailable
(Section 3.2.4). Stimulation of the microbial community may or may not have
concurrent effects on soil invertebrates depending on whether the increase in
microbial biomass is exposed for predation. If the majority of the increase in
microbial biomass occurs within biochar particles in the soil, then the
microorganisms may not be available as a food source for soil invertebrates.
However, if the stimulated growth in microbial biomass also occurs outside of
biochar particles within the soil, then it is possible that an increase in the soil
invertebrate community may also occur. This could have implications for
nutrient cycling, crop yield and other ecosystem services which are hard to
predict owing to a paucity of experimental data and the high intrinsic
complexity and dynamic nature of the edaphic community.

3.2.6.3 Soil megafauna

There is no research reported in the literature on the effects of charcoal or
biochar addition to soil on soil megafauna such as badgers, moles or other
vertebrates. As these organisms are not generally found in the arable
environment it is likely that any effects may be minimal if biochar addition is
limited to agricultural land. However, should biochar addition be planned for
other soils, including forest soils, then an impact assessment may well need
to be carried out to investigate any possible impacts.
Off-site effects of biochar addition to arable soils are possible, and are likely to
include any contaminants such as heavy metals moving up through the food
chain. This is likely to be particularly true of moles that have a diet high in
earthworms. As it has been shown that earthworms ingest charcoal which
exists within the soil profile, it is probable that moles will in turn ingest
charcoal particles when they ingest worms. It is still currently unclear what
quantity of heavy metals, if any, will be able to pass from the biochar, if
present, into the tissues of other organisms and this is an area which requires
significant further work to ensure the safety of heavy metal containing
biochars in soils (see also Section 3.2.4).
The main point of contact between biochar in the soil and other megafauna
such as rabbits, badgers and foxes is likely to be through skin contact when
the animals are building and resting in their burrows, sets and ‘earths’. Heavy
metal absorption is extremely limited through skin, with the exception of
mercury which is likely to exist in biochars in extremely minute amounts, if at
all. It is possible that some small amount of biochar may enter these
organisms’ digestive tracts and airways if it is in the form of very small
particles, as well as through ingestion of earthworms in the case of some

90
organisms such as moles, as earthworms have been shown to ingests
charcoal in soil (Topoliantz and Ponge, 2005).
Concerning possible ingestion of biochar fragments from the soil by soil
megafauna there is no published data in the primary literature. However, Van
et al. (2006) found that incorporation of bamboo charcoal (0.5 to 1.0 mg kg-1
of body weight) into the feed of growing goats resulted in enhanced growth
and no adverse effects were observed at the study concentrations. Clearly
care must be taken when extrapolating data to other animals and to biochars
made from alternative feedstocks and this area warrents further research.
Ingestion is not the only mode of possible uptake of biochar fragments by the
soil megafauna. Biochar dust particles may possibly be inhaled by the soil
megafauna. However, there is currently no research reported in the primary
literature concerning the effect of charcoals on the respiratory systems of soil
megafauna and as such robust predictions concerning the possible effects is
currently not possible and requires further research.

3.3 Production Function

Increased yields are the most commonly reported benefits of adding biochar
to soils. Nearly all experiments have been conducted in the tropics, while field
trials in temperate regions have been set up only recently. Taking a step back,
SOM is generally believed to be correlated positively with crop yields in
modern arable agriculture, although there is still poor scientific understanding
of the strength of this relationship, the influence of environmental conditions
(sandy or clayey soil, wet, dry, etc.), crop types (combinable vs. root crops)
and the underlying mechanisms. Loveland and Webb (2003) reviewed 1200
papers in the scientific literature on the relationship between SOC and crop
yield in temperate regions and concluded that a consensus does not exist.
Diaz-Zorita et al. (1999) performed stepwise regression analysis between
wheat yields and soil properties and found different relationships in different
years. In a year without a water deficit, N and P influenced yield, in drought
years however, yields were correlated to water availability and OM. Pan and
Smith (2009) investigated the relationship between SOM and yield by using
statistical data for China (1949-1998) and found a particularly strong
relationship between yield stability and SOM.
Considering the poor understanding of the relationship between SOM and
crop yield or plant production, it may be expected that similar challenges exist
regarding the scientific understanding of the relationship between biochar and
plant productivity. However, to investigate the relationship between biochar
additions to soils and crop productivity in more detail, new tools can be used.
Therefore, a meta-analysis on this relationship was performed (see next
sections).

3.3.1 Meta-analysis methods

Objectivity of systematic reviews on biochar is paramount. In the medical
sciences this has been resolved by the founding of an independent
organisation (the Cochrane Collaboration) that provides regularly updated
systematic reviews on specific healthcare issues using a global network of
volunteers and a central database/library. The methodologies used in medical

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science can be transferred to ensure objectivity when compiling literature
reviews in other research areas such as those related to biochar, even though
the amount of literature and information available on biochar is currently
limited. One such methodology which was developed for objective analysis of
a range of different medical studies testing the same (or similar) hypothesis
was that of meta-analysis which is being increasingly used across a range of
scientific disciplines.
Here, meta-analysis techniques (Rosenberg et al., 1997) were used to
quantify the effect of biochar addition to soil on plant productivity. For each
study the control mean and experimental means were recorded, or calculated
where necessary. Standard deviation was used as a measure of variance and
this was reported where given or calculated from the published measure of
variance from each study. To maximise the number of studies used in the
analysis, both pot and field experiments were recorded, providing the results
were quantitative.
Standardisation of the results from the studies was undertaken through
calculation of the “effect size” which allows quantitative statistical information
to be pooled from and robust comparisons of effects from studies with
different variables to be made. Data was square root transformed to normalise
the distribution. Effect size was calculated using the transformed data taken
as the natural logarithm of the response ratio by using the following equation:
⎛ E⎞
ln R = ln⎜ x C ⎟
⎜ ⎟
⎝x ⎠
E C
Where x = mean of experimental group; and x = mean of control group

For the meta-analysis, the following nine studies concerning the effects of
biochar addition to soil on crop production were used: Van Zwieten et al.
(2008); Yamato (2006); Chan (2007); Chan (2008); Lehmann (2003); Ishii and
Kadoya (1994); Nehls (2002); Kimetu et al. (2008) and Blackwell (2007).
These studies combined produced 86 different ‘treatments’ for use in the
meta-analysis.
In order to use change in pH as a grouping category, changes were grouped
by ‘no change’ (0 – representing no change from soil starting pH upon
addition of biochar) and in consecutive changes in pH of 0.5 for both
increasing and decreasing pH values upon biochar addition. For calculation of
grouped effect sizes a categorical random effects model was used. Groups
with fewer than two variables were excluded from each analysis. Resampling
tests were generated from 999 iterations. For each of the analyses, grouped
by different categorical predictors, the data was analysed using a fixed effects
model if the estimate of the pooled variance was less than or equal to zero.
When plotting figures, the effect size was unlogged (exponentially
transformed) and the result multiplied by 100 to obtain the percentage change
in effect size upon biochar addition in each category. Analysis was
undertaken using MetaWin Version 2 statistical software (Rosenberg et al.,
2000). While more than the nine reported studies looking at the effect of

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biochar addition to soil on crop productivity, studies were excluded from the
analysis when no quantitative results or measures or variance were available,
leaving the nine studies reported above.

3.3.2 Meta-analysis results

Figure 3.2 The percentage change in crop productivity upon application of biochar at different
rates, from a range of feedstocks along with varying fertiliserco-amendments. Points represent
mean and bars represent 95% confidence intervals. Numbers next to bars denote biochar
application rates (t ha-1). Numbers in the two columns on the right show number of total
‘replicates’ upon which the statistical analysis is based (bold) and the number of ‘experimental
treatments’ which have been grouped for each analysis (italics)

Figure 3.2 shows the effect of biochar addition to soil on crop productivity,
grouped by application rate and vertically partitioned by effect size. The
sample means seem to indicate a small but positive effect on crop productivity
with a grand mean (being the mean of all effect sizes combined) of about
10%. There appears to be a general trend, when looking at the sample
means, for increased biochar application rate to be correlated with increased
crop productivity (Figure 3.2). However, there was no statistically significant
difference (at P = 0.05) between any of the application rates as is evident from
the overlapping error bars which represent the 95% confidence intervals.
Application rates of 10, 25, 50 and 100 t ha-1 were all found to significantly
increase crop productivity when compared to controls which received no
biochar addition. However, other application rates which fall within the range
of these statistically significant application rates, such as 40 and 65 t ha-1
showed no statistically significant effect of biochar addition to soil on crop
yield, demonstrating that while biochar addition to soil may increase crop
productivity it is not linearly correlated.
It can be seen from Figure 3.2 that even with the same application rate of
biochar, a large variation in effect size occurs. This is particularly true of the

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lower application rates of 5.5 and 11 t ha-1 and also for the large application
rate of 135.2 t ha-1. Other application rates also have a large variance in their
effect size, but to a lesser extent. The reason for this large variation is likely to
be due to the different biochar feedstocks used, the different crops assessed
and differences in soil type to which the biochar was added. It is interesting to
note that while there was often large variation in the data for a given
application rate, the means for each application rate all fall on the positive
productivity effect side, and no single biochar application rate was found to
have a statistically significant negative effect on the crops from the range of
soils, feedstocks and application rates studied. It should be noted that while
no negative effects have been detected by this meta-analysis with regard to
the effect of application rate on crop productivity, the studies used in the
meta-analysis do not cover a wide range of latitudes and the data used was
heavily scewed towards (sub)tropical conditions. This means that while this
analysis provides good evidence of the generally positive effects of biochar
addition to soil on crop productivity, care needs to be taken when
extrapolating these results to European latitudes, crops and soil types.

Figure 3.3 Percentage change in crop productivity upon application of biochar at different rates
along with varying fertiliserco-amendments grouped by change in pH caused by biochar addition
to soil. Points represent mean and bars represent 95% confidence intervals. Values next to bars
denote change in pH value. Numbers in the two columns on the right show number of total
‘replicates’ upon which the statistical analysis is based (bold) and the number of ‘experimental
treatments’ which have been grouped for each analysis (italics)

Figure 3.3 shows the effect of biochar addition to soil on crop productivity,
grouped by liming effect. It should be noted that where the biochar addition to
soil lead to a liming effect (i.e. the pH of the soil was increased), there was a
significant increase in crop productivity compared to controls, although there
were no significant differences between treatments which lead to a positive
liming effect.

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Regarding those treatments that showed no change, or a reduction in pH
upon biochar addition to soil, biochar addition to soil showed no statistically
significant effect. All other groupings where biochar addition to soil led to an
increase in soil pH, a concurrent increase in crop productivity was seen. This
effect was not strictly linear, with the mean increase in crop productivity where
biochar caused a liming effect (with an increase in pH units ranging from 1.1-
1.5), was lower when compared to those treatments where the liming effect
resulted in an increase ranging from 0.6 to 1.0 pH units. This may be due to
differences in initial pH, before biochar addition to soil, meaning that a lesser
increase was still sufficient to pass a tipping point with regard to metal ion
availability for example, meaning a slightly increased crop productivity effect
even with a decreased liming effect.

Figure 3.4 The percentage change in crop productivity o upon application of biochar at different
rates along with varying fertiliserco-amendments to a range of different soils. Points shows mean
and bars so 95% confidence intervals. Numbers in the two columns on the right show number of
total ‘replicates’ upon which the statistical analysis is based (bold) and the number of
‘experimental treatments’ which have been grouped for each analysis (italics)

Figure 3.4 shows the effect of biochar addition to soil on crop productivity,
grouped by soil type. As with the previous meta-analysis figures, the error
bars are again very large. Again, there were found to be no statistically
significant negative effects of biochar to soil on crop productivity when
grouped by soil type. The trend of the effect in Calcarosols was towards the
negative, but this effect was not statistically significant when compared to
control soils, although it was significantly less than then positive effects seen
upon biochar addition to both loam soils and acidic free draining soils. The
effect of biochar addition to these soils (‘loam’ and ‘acidic free draining’) was
also found to show a statistically significant increase when compared to
control soils with no biochar addition. For the other soil types investigated by
this analysis (‘volcanic parent material’ and ‘free draining’), there was a
general trend towards a positive effect as evidenced by the means being on

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the positive effect side of 0. However, the effect for these soils was not found
to be statistically significant owing to the large variation from the samples.

Figure 3.5 The percentage change in crop productivity of either the biomass or the grain upon
application of biochar at different rates along with varying fertiliserco-amendments. Points
shows mean and bars so 95% confidence intervals. Numbers in the two columns on the right
show number of total ‘replicates’ upon which the statistical analysis is based (bold) and the
number of ‘experimental treatments’ which have been grouped for each analysis (italics)

Figure 3.5 shows the effect of biochar addition to soil on crop productivity,
grouped by overall biomass productivity vs. grain yield. There was no
significant difference in grain yield for those crops grown in biochar amended
soils compared to non-biochar amended soils. There was a significant
increase in overall crop biomass production in biochar amended soils
compared to non-biochar amended soils, although this difference was not
significant when compared to the impact of growth on biochar amended soils
on grain production.
The fact that biomass was positively affected by growth on biochar amended
soils whereas grain was not is possibly due to grain being a relatively small
part of the biomass and so any slight change would be more difficult to detect.
Again, the error bars show that there was considerable variation within
treatments, as would be expected due to the data being amalgamated from
several different studies, and each treatment in the above figure includes data
obtained from different crops, soils and biochar feedstocks.

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Figure 3.6 The percentage change in crop productivity upon application of biochar along with a
co-amendment of organic fertiliser(o), inorganic fertiliser(I) or no fertiliser(none). Points shows
mean and bars so 95% confidence intervals. Numbers in the two columns on the right show
number of total ‘replicates’ upon which the statistical analysis is based (bold) and the number of
‘experimental treatments’ which have been grouped for each analysis (italics)

There was no statistically significant difference between biochar application to

soil whether no concurrent fertiliser addition was used, or whether organic or
inorganic fertiliser was used (Figure 3.6). This is contrary to what is often
reported in the literature where specific recommendations often state that
fertiliser addition is necessary to maximise crop yields.
Care must be taken when interpreting Figure 3.6, as it appears at first glance
to show no difference in effect size between addition of biochar alone, or with
fertiliser. It is important to remember that the effect sizes are between
‘controls without biochar’ vs ‘treatments with biochar’. This means that the no
fertiliser application treatment shows the effect of biochar addition to soil
alone. In the other treatments, the control includes the addition of fertiliser, but
without the addition of biochar, compared to the experimental treatments
which include both fertiliser and biochar. Figure 3.6 shows, therefore, that the
impact of biochar addition to soil was not significiantly different whether
fertiliser, either organic or inorganic was used. This does not show, as
appears at first glance, that there was no significant effect of co-addition of
fertiliser with biochar, over addition of biochar to soil alone.
While there was found to be no significant difference between the effects of
inorganic fertiliser with biochar compared to no fertiliser with biochar, both of
these treatments showed increased crop productivity when compared to
control non-biochar amended soils. Chan et al. (2007) reported a lack of
response upon addition of biochar without the co-addition of N and as such it
seems likely that in those studies available N in the soil was not a limiting
factor, possibly due to previous cropping with legumes, or owing to the
quantity and quality of SOM meaning that available N levels were not limiting.

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The addition of organic fertiliser along with biochar to soils was found to have
no statistically significant effect when compared to application of organic
fertiliser to non-biochar amended soils. This is due to extreme levels of
variance in results of the biochar plus organic fertiliser treatments, as shown
by the large error bars.

3.3.3 Meta-analysis recommendations

As was shown in this report, soils are very heterogeneous systems, in both
time and space and at a multitude of scales, and biochar is a very
heterogeneous material. Meta-analysis is a valuable tool for amalgamating,
summarising and reviewing studies on biochar. It can elucidate trends in a
quantitative way that in conventional reviews might be perceived as being
biased by personal judgement. A combination of meta-analysis with a
qualitative review of the literature will provide the most comprehensive
discussion of both the status of scientific knowledge on a specific ‘effect’ and
the possible underlying mechanisms and exceptional or marginal conditions.
As new studies are published, the meta-analyses on the effect of biochar
application to soil on productivity can be updated (and refined) periodically. In
addition, many other effects of biochar (see Chapter 3) can be analysed by
meta-analyses once a large enough body of research has been established.
From this work it is strongly recommended that scientists publishing results on
effects of biochar describe the data, and the variance of those data,
consistently and completely. This means including the Z or F statistic for
regression data and clear measures of variance for comparative analysis
data, such as standard deviations or standard errors for each treatment,
including the control, rather than an LSD (least significant difference) which
has been pooled for several treatments. In all cases, it should be absolutely
clear what the sample number is for every treatment (including control).
Clearly this should be normal scientific conduce, but unfortunately does not
seem to occur in all cases. To enable meta-analyses on effects of a factor that
is not the dependent variable of a study, it is also recommended to include all
sample numbers, standard deviations or standard errors of other parameters
measured in the study, e.g. CEC, pH, bulk density, microbial activity, etc.
Finally, it is recommended to report all the data in tabular format, possibly as
an annex.

3.3.4 Other components of crop production function

Crop production is, however, only one possible agronomic effect of on-farm
benefit from biochar. Many other effects still need to be investigated, for
example i) direct impacts on yields (seed rate); ii) crop-related impacts (crop
establishment, fertiliser, disease and weeds); and iii) non-crop-related impacts
(workability, soil hydrology, soil degradation).

3.4 Summary
This section has highlighted the relative paucity of knowledge concerning the
specific mechanisms behind the reported interactions of biochar within the soil
environment. However, while there is still much that is unknown, large steps
have been taken towards increasing our understanding of the effects of
biochar on soil properties and processed. Biochar interacts with the soil

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system on a number of levels. Sub-molecular interactions with clay and silt
particles and SOM occur through Van der Waals forces and hydrophobic
interactions. It is the interactions at this scale which will determine the
influence of biochar on soil water repellency and also the interactions with
cations and anions and other organic compounds in soil. These interactions
are very char specific, with the exact properties being influenced by both the
feedstock and the pyrolysis conditions used.
There has been some evidence to suggest that biochar addition to soil may
lead to loss of SOM via a priming effect in the short term. However, there is
only very little research reported in the literature on this subject, and as such it
is a highly pertinent area for further research. The fact that Terra Pretas
contain SOM as well as char fragments seems to demonstrate that the
priming effect either does not exist in all situations or if it does, perhaps it only
lasts a few seasons and it appear not to be sufficient to drive the loss of all
native SOM from the soil. Biochar has the potential to be highly persistent in
the soil environment, as evidenced both by its presence in Terra Pretas, even
after millennia, and also as evidenced by studies discussed in this section.
While biochars are highly heterogeneous across scales, it seems likely that
properties such as recalcitrance and effects on water holding capacity are
likely to persist across a range of biochar types. It also seems probable, that
while difference may occur within biochars on a microscale, biochars
produced from the same feedstocks, under the same pyrolysis conditions are
likely to be broadly similar, with predictable effects upon application to soil.
What remains to be done are controlled experiments with different biochars
added to a range of soils under different environmental conditions and the
precise properties and effects identified. This may lead towards biochars
possibly being engineered for specific soils and climate where specific effects
are required.
After its initial application to soil, biochar can function to stimulate the edaphic
microflora and fauna due to various substrates, such as sugars, which can be
present on the biochar's surface. Once these are metabolised, biochar
functions more as a mineral component of the soil rather than an organic
component, as evidenced by its high levels of recalcitrance meaning that it is
not used as a carbon source for respiration. Rather, the biochar functions as a
highly porous network the edaphic biota can colonise. Due to the large
inherent porosity, biochar particles in soil can provide refugia for
microorganisms whereby they may often be protected from grazing by other
soil organisms which may be too large to enter the pores. This is likely to be
one of the main mechanisms by which biochar-amended soils are able to
harbour a larger microbial biomass when compared to non-biochar amended
soils. Biochar incorporation into soil is also expected to enhance overall
sorption capacity of soils towards trace anthropogenic organic contaminants
(e.g. PAHs, pesticides, herbicides), in a stronger way, and mechanistically
different, from that of native organic matter. Whereas this behaviour may
greatly contribute to mitigating toxicity and transport of common pollutants in
soil, biochar aging over time may result in leaching and increased
bioavailability of such compounds. On the other hand, while the feasibility for
reducing mobility of trace contaminants in soil might be beneficial, it might

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also result in their localised accumulation, although the extent and
implications of this have not been experimentally assessed.
Soil quality may not be necessarily improved by adding biochar to soil. Soil
quality can be considered to be relatively high for supporting plant production
and provision of ecosystem services if it contains carbon in the form of
complex and dynamic substances such as humus and SOM. If crop residues
are used for biochar, the proportion of carbon going into the dynamic SOM
pool is likely to be reduced, with the carbon being returned to the soil in a
relatively passive biochar form. The proportion of residues which are removed
for pyrolysis versus the proportion which is allowed to remain in the soil will
determine the balance between the dynamic SOM and the passive biochar
and so is likely to affect soil quality for providing the desired roles, be it
provision of good use as crop or timber, or functioning as a carbon pool.
Biochar also has the potential to introduce a wide range of hazardous organic
compounds (e.g. heavy metals, PAHs) into the soil system, which can be
present as contaminants in biochar that has been produced either from
contaminated fedstocks or under processing conditions which favour their
production. While a tight control over the feedstock type and processing
conditions used can reduce the potential risk for soil contamination,
experimental evidence of the occurrence and bioavailability and toxicity of
such contaminants in biochar and biochar-enriched soil (over time) remain
scarce. A comprehensive risk assessment of each biochar product prior to its
incoporation into soil, taking into account the soil type and environmental
conditions, is therefore paramount.
Increased crop yields are the most commonly reported benefits of adding
biochar to soils. A full search of the scientific literature led to a compilation of
studies used for a meta-analysis of the effects of biochar application to soils
and plant productivity. Meta-analysis techniques (Rosenberg et al., 1997)
were used to quantify the effect of biochar addition to soil on plant productivity
from a range of experiments. Our results showed a small overall, but
statistically significant, positive effect of biochar application to soils on plant
productivity in the majority of cases, covering a range of both soil and crop
types. The greatest positive effects were seen on acidic free-draining soils
with other soil types, specifically Calcarosols showing no significant effect. No
statistically significant negative effects were found. There was also a general
trend for concurrent increases in crop productivity with increases in pH up on
biochar addition to soils. This suggests that one of the main mechanisms
behind the reported positive effects of biochar application to soils on plant
productivity may be a liming effect. These results underline the importance of
testing each biochar material under representative conditions (i.e. soil-
environment-climate-management factors).
The degree and possible consequences of the changes biochar undergo in
soil over time remain largely unknown. Biochar loss and mobility through the
soil profile and into water resources has so far been scarcely quantified and
the underlying transport mechanisms are poorly understood. This is further
complicated by the limited amount of long-term studies and the lack of
standardized methods for simulating biochar aging and for long-term
environmental monitoring.

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4. BIOCHAR AND ‘THREATS TO SOIL’
This chapter summarises the findings and gaps in the biochar literature
relevant to the threats to soil, as identified by the Thematic Strategy for Soil
Protection (COM(2006) 231). For a more in-depth discussion of patterns,
effects, processes and mechanisms, please refer to the relevant Sections in
this report. For the threats to soil of ‘soil sealing’ and ‘landslides’, biochar
holds no relevance at present.

4.1 Soil loss by erosion

In the context of threats to soil, soil loss by erosion is specified by being “as a
result of anthropogenic activity, in excess of natural soil formation rates
causing a deterioration or loss of one or more soil functions” (Jones et al.,
2008). Experimental studies on the effects of biochar application on soil
erosion have not been found. Even erosion of charcoal particles from the soil
surface after wildfires is a topic that has only started being researched
relatively recently. However, an obvious potential effect is the wind erosion of
biochar particles during application to soils. For application strategies where
the biochar is incorporated into the soil, further erosion by either wind is likely
to be reduced to the ‘normal’ erosion rates of the site. For application
strategies where the biochar is applied to the soil surface only, the risk of
erosion increases strongly because biochar generally has a relatively low
density and, therefore, a greater erodibility by wind for smaller particles and
by water for also the larger biochar particles. Surface application has been
discussed for grassland and forest land uses mostly (and no-till systems). The
greater risk may be expected for grasslands since these are open systems
with generally greater wind velocities than forests.
Biochar application to soils can also be considered from a soil formation
perspective. Verheijen et al. (2009) reviewed soil formation rates in Europe to
be in the range of 0.3-1.4 t ha-1 yr-1. Considering the human life span, these
very low formation rates (measurable only in geological terms) mean that soil
is a non-renewable resource. Even low application rates of biochar are likely
to outstrip natural soil formation rates by physicochemical weathering and
dust deposition (i.e. mineral dust mainly from the Sahara). However, great
care must be taken when considering biochar application to soils as
constituting towards soil formation rates, and thereby tolerable soil erosion
rates. Most notably, the residence time of biochar particles in soils needs to
be considered, which depends on i) decomposition rates of biochar
components (physicochemical and biological degradation), and ii) mobility and
fate of biochar particles (movement through the soil matrix and into
ground/surface waters). Both these factors are likely to be influenced strongly
by variation in soil properties, climatic conditions, biochar properties, and land
use and soil management. A substantial body of experimental scientific
research into the mechanisms affecting the residence time of biochar particles
in soils is required before biochar application to soils might be considered in
the context of tolerable soil erosion rates. Conventionally, SOM build up is not
considered for soil formation rates of mineral soils. Under what conditions
those components of biochar that are very recalcitrant (e.g. residence time
>1,000 yr) will reside in the soil matrix during their ‘life span’, is unknown at
present. The interaction between biochar particles, mineral soil particles and

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native organic matter (NOM), or OM that is applied with (or after) the biochar,
is likely to play a major role (see Section 3.2.1 and 3.2.5).
Wind erosion is caused by the simultaneous occurrence of three conditions:
high wind velocity; susceptible surface of loose particles; and insufficient
surface protection. Theoretically, if biochar particles are produced with water
retention properties greater than the water retention capacity of the soil
surface at a site, and if the biochar particles become a structural component
of that surface soil (e.g. not residing on top of the soil surface), and possibly
interacting with OM and mineral particles, then wind erosion rates at that site
may be reduced, all other factors remaining equal. The application of biochar
dust to the soil surface (i.e. not incorporated) can pose risks via wind erosion
of the dust particles and subsequent inhalation by people. Strict guidelines on
biochar application strategies under specific environmental and land use
conditions could prove sufficient to prevent this risk.
Water erosion takes place through rill and/or inter-rill (sheet) erosion, and
gullies, as a result of excess surface runoff, notably when flow shear stresses
exceed the shear strength of the soil (Kirkby et al., 2000, 2004; Jones et al.,
2004). This form of erosion is generally estimated to be the most extensive
form of erosion occurring in Europe. If biochar reduces surface runoff, then,
logically, it will reduce soil loss by water erosion, all other factors remaining
equal. Surface runoff can be reduced by increased water holding capacity
(decreasing saturation overland flow) or increased infiltration capacity
(decreasing infiltration excess – or Hortonian - overland flow) of the topsoil.
Under specific environmental conditions, it seems that biochar with large
water retention properties could diminish the occurrence of saturation
overland flow. This effect could be enhanced when biochar addition leads to
stabilisation of NOM, or OM that is added with, or after, the biochar. Infiltration
excess overland flow depends more on soil structure and related drainage
properties. In particular the soil surface properties are important for this
mechanism. It is not inconceivable that specific biochar particles can play a
role in increasing infiltration rates, however, other biochar particles could also
lead to reduced infiltration rates when fine biochar particles fill in small pore
spaces in topsoils, or increased hydrophobicity (Section 3.1). In addition, and
this could be an overriding factors at least in the short term, the biochar
application strategy and timing is a potential source of topsoil and/or subsoil
compaction (Section 1.8) and, thereby, reduced infiltration rates.
It stands to reason that under those conditions where surface runoff is
reduced by biochar application, possibly as part of a wider package of soil
conservation measures, a concomitant reduction in flooding occurrence and
severity may be expected, all other factors remaining equal. However, as
stated at the beginning of this section, experimental evidence of biochar
application on erosion was not encountered in the scientific literature, nor was
it for flooding. On the other hand, under conditions where biochar application
leads to soil compaction (see Section 1.8) runoff may be increased leading to
more erosion. Research is needed into all aspects of effects from biochar
addition to soil loss by erosion described here, and in particular into the
mechanisms behind the effects. Even a small effect may be worthwhile
considering estimates of the cost to society from erosion. For example annual
costs have been estimated to be £205 million in England and Wales alone

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and $44 billion in the U.S.A. (Pimentel et al., 1995). In addition, active and
targeted modification of the water retention function of specific soils could be
considered in the context of scenarios of adapting to changing rainfall patterns
(seasonal distribution, intensity) with climate change. In the future, climate
change looks likely to increase rainfall intensity over large areas of Europe, if
not annual totals, thereby increasing soil erosion by water, although there is
much uncertainty about the spatio-temporal structure of this change as well as
the socio-economic and agronomic changes that may accompany them (e.g.
Boardman and Favismortlock, 1993; Phillips et al.,1993; Nearing et al., 2004).

4.2 Decline in soil organic matter

Decline in SOM is defined as a negative imbalance between the build-up of
SOM and rates of decomposition leading to an overall decline in SOM
contents and/or quality, causing a deterioration or loss of one or more soil
functions (Jones et al., 2008).
The interaction between biochar and NOM, or OM that is added with the
biochar, or afterwards, is complex. Many mechanisms have been identified
and are discussed in this report, i.e. priming effect, residue removal, liming
effect, organomineral interactions, aggregation and accessibility.

Biochar replacing peat extraction

If biochar is engineered to have good plant-available water properties as well as

nutrient retention, it could come to replace peat as a growing medium in
horticulture (also agriculture), and as a gardening amendment sold in garden
centres. Peatlands currently used (‘mined’) for peat extraction could then be
restored with substantial benefits to their functioning and the ecosystem services
which they provide, e.g. maintenance of biodiversity, C sequestration, water
storage, etc. Janssens et al. (2005) reported that undisturbed European
peatlands sequester C at a rate of 6 g m-2 total land area, while peat extraction
caused a C loss of 0-36 g m-2 total land area. Janssens et al. (2003) estimated a
net loss of 50 (±10) Mt yr-1 for the European continent, which is equivalent to
around 1/6 of the total yearly C loss from European croplands. However, this
value is likely to be greater when also considering C emissions associated with
continued decomposition at abandoned peat mines (Turetsky et al., 2002),
transport to processing plant, transport to market, and decomposition of the
applied peat (e.g. in a life cycle assessment; Cleary et al., 2005).

4.3 Soil contamination

Recently, increasing knowledge on the sorption capacity of biochar has had
two environmentally important outcomes. Firstly, the realisation that biochar
addition to a soil can be expected to improve its overall sorption capacity, and
consequently influence the toxicity, transport and fate of any organic
compounds, which may be already present or are to be added to that soil (see
Section 3.2.2). Secondly, enhanced awareness that biochar from widely
available biomass resources can be applied to soils and sediments as a low-

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cost and low-environmental-impact mitigation/remediation strategy for
common environmental pollutants.
The latter outcome appears to be even more attractive when considering the
time and cost benefits associated to biochar production, relatively to that of
activated carbon in various applications. Activated carbon results from
activating (involving partial oxidation) a charcoal precursor by means of
exposing it to CO2, steam or acid at high temperatures, in order to further
increase its surface area (per gram; McHenry, 2009). Overall, evidence
suggests that biochar and activated carbon have comparable sorption
affinities, as demonstrated by Tsui and Roy (2008), using compost biochar
(pyrolysis temperatures ranging between 120-420°C) and corn stillage
activated carbon for removal of the herbicide atrazine in solution (1.7 mg L-1).
In fact, the effectiveness of activated carbon over that of wood biochar has
been questioned in some instances (Pulido et al., 1998; Wingate et al., 2009),
but this aspect remains far from fully evaluated.
Wingate et al. (2009) have very recently patented the development and
application of charcoal from various plant and crop tissues (leaves, bark and
stems) of ammonium (NH4+) and heavy metal-contaminated environments
(soil, brown-field site, mine tailings, slurry, and aqueous solution). Heavy
metal ions are strongly adsorbed onto specific active sites containing acidic
carboxyl groups at the surface of the charcoal (e.g. Machida et al., 2005).
Surprisingly, the mechanism of metal uptake by charcoals appears to involve
replacing pre-existing ions contained in the charcoal (e.g. K, Ca, Mg, Mn,
excluding Si), with the metal ion, suggesting a relationship between the
mineral content of the charcoal and its remediation potential for heavy metals
(Wingate et al., 2009).
In the soil environment, biochar has already been shown to be effective in
mitigating mobility and toxicity of heavy metals (Wingate et al., 2009) and
endocrine disruptors (Smernik, 2007; Winsley, 2007). However, very little
work of this kind has been accomplished and data is still scarce. It is likely
that soil heterogeneity and the lack of monitoring techniques for biochar in this
environment may partly explain such a gap. The previous discussion on
contaminant leaching over time as a consequence of biochar aging in the
environment (see Section 3.2.1) does not necessary mean that its high
remediating potential should be disregarded. For example, it could be
employed as a ‘first-instance’ pollutant immobilisation from point sources.
Also, biochar’s higly porous matrix might be ideal as carrier for microrganisms
as part of bioaugmentation programs for specific sites, where indigenous
microbial populations are scarce or have been suppressed by the
contaminant (Wingate et al., 2009). In this context, for instance, Wingate et al.
(2009) have reported the successful application of charcoal carrying 1010
hydrocarbon degraders (per gram of charcoal) in diesel-polluted sites,
resulting in 10 fold enhancement of hydrocarbon degradation in this
environment. Clearly, it is likely that appropriate regulatory requirements for
cleanup and closure would be needed before any remediation plan involving
biochar could be implemented. Experimental evidence is required in order to
verify this.

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There is also evidence that it is possible to use biochar’s sorptive capacity in
water and wastewater treatments (Wingate et al., 2009), whereas the use of
activated carbon for removal of chlorine and halogenated hydrocarbons,
organic compounds (e.g. phenols, PCBs, pesticides) and heavy metals
(Boateng, 2007) has long been established. Crop residue (mainly wheat)
biochar produced at temperatures between 300°C and 700°C has already
shown potential for removal of sulphate (Beaton, 1960), benzene and
nitrobenzene from solution (Chun et al., 2004), while bamboo charcoal
powder has been effective in uptake of nitrate from drinking water (Mizuta et
al., 2004). Other studies in aqueous media have reported biochar’s capacity
to adsorb phosphate and ammonium (Lehmann et al., 2002; Lehmann et al.,
2003, 2003b), with further applications having been reviewed by Radovic et
al. (2001). In the context of water treatment, Sohi et al. (2009) have pointed
out that a higher control over the remediation process would be achievable,
comparatively to that in soil.
The possibility of using ‘engineered’ (or ‘tailor-made’) biochar (Pastor-Villegas
et al., 2006) in order to meet the requirements for a specific remediation plan
looks increasingly promising. As the mechanisms of biochar production,
behaviour and fate, as well as its impact on ecosystem health and functioning
become increasingly well understood, biochar can be optimised to deliver
specific benefits (Sohi et al., 2009). Nevertheless, data on competitive
sorption in soils and sediments emphasize the need for a full characterisation
of the contaminated site and the coexisting chemical species before any
remediation plan involving biochar is put in place.

4.4 Decline in soil biodiversity

Decline is soil biodiversity is defined as a ‘reduction of forms of life living in the
soil (both in terms of quantity and variety) and of related functions, causing a
deterioration or loss of one or more soil functions‘ (Jones et al., 2008). There
is evidence of decline in soil biodiversity in some specific cases. For example,
the Swiss Federal Environment Office has published the first-ever “Red List”
of mushrooms detailing 937 known species facing possible extinction in the
country (Swissinfo 2007). In another instance, the New Zealand flatworm is
increasing in numbers and extent and potentially poses a great threat to
earthworm diversity in the UK with a 12% reduction in earthworm populations
in some field sites in Scotland already reported (Boag et al. 1999). Changes in
earthworm community structure have been also recorded (Jones et al., 2001).
The exact impacts of a decline in soil biodiversity are far from clear, due to
complications by such phenomena as functional redundancy. However, it is
clear that any decline in soil biodiversity has the potential to compromise
ecosystem services, or at least reduce the resistance of the soil biota to
further pertubations. Although evidence exists for declines in soil biodiversity
in some specific cases, it is a highly depauperate area of research. However,
no studies have been published to date looking at how biochar additions to
soil can be used to restore soil biodiversity to previous levels in any given
area.
Threats to soil biodiversity consist of those soil threats as described in the
Thematic Strategy for Soil Protection (COM(2006) 231) and as such, in those

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situations where biochar either helps the mitigation of, or increases the
problem of, it is likely that knock on effects for the soil biota will occur.

4.6 Soil compaction

Soil compaction is defined as the densification and distortion of soil by which
total and air-filled porosity are reduced, causing a deterioration or loss of one
or more soil functions (Jones et al., 2008).
The effects of biochar on soil compaction have been studied very little. Both
potential positive and negative effects may occur, for topsoil as well as subsoil
compaction. Whereas topsoil compaction is ‘instantaneous’, subsoil
compaction is a cumulative process leading to densification just below the
topsoil over the years. A biochar application strategy, where application
occurs every year, is, therefore, a greater risk of subsoil compaction than a
‘single application’ biochar strategy. An obvious risk of compaction is the
actual application of biochar itself. When applied with heavy machinery and
while the water-filled pore volume of soil is high, the risk of compaction
increases. Biochar also has a low elasticity, measured by the relaxation ratio
(R), which is defined as the ratio of the bulk density of the test material under
specified stress to the bulk density after the stress has been removed. Straw
has a very high elasticity ratio and, therefore, when straw is charred and
applied as biochar instead of fresh straw, the resilience of the soil to
compactive loads is reduced, all other factors remaining equal. The bulk
density of biochar is low and, therefore, adding biochar to soil can lower the
bulk density of the soil thereby reducing compaction. However, when biochar
is applied as very fine particles, or when larger biochar particles disintegrate
in arable soils under influence of tillage and cultivation operations, these can
fill up small pores in the soil leading to compaction.
Compaction by machinery may be prevented relatively easily by promoting
sound soil management. However, compaction by the behaviour of biochar
particles in the soil has received very little attention in research so far and
mechanisms are understood poorly.

4.7 Soil salinisation

Soil salinisation is defined as the accumulation of water soluble salts in the
soil, causing a deterioration or loss of one or more soil functions. The
accumulated salts include sodium-, potassium-, magnesium- and calcium-
chlorides, sulphates, carbonates and bicarbonates (Jones et al., 2008). A
distinction can be made between primary and secondary salinisation
processes. Primary salinisation involves accumulation of salts through natural
processes as physical or chemical weathering and transport processes from
salty geological deposits or groundwater. Secondary salinisation is caused by
human interventions such as inappropriate irrigation practices, use of salt-rich
irrigation water and/or poor drainage conditions (Huber et al., 2009). Salts
associated with biochar should be considered as a potential source for
secondary salinisation.
Various salts can be found in the ash fraction of biochar, depending mostly on
the mineral content of the feedstock. Indications are that the ash content of
biochar varies from 0.5% - 55%. In classic charcoal manufacturing, ‘good

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quality’ charcoal is referred to as having 0.5% – 5.0% ash (Antal and Gronli,
2003). However, biochar produced from feedstocks such as switchgrass and
maize residue have been reported to have an ash content 26% - 54% much of
which as silica, while hardwood ash contains mainly alkali metals (Brewer et
al., 2009). A wide range of trace elements have been measured in biochar
ash, e.g. boron, cupper, zinc, etc., however, the most common elements are
potassium, calcium, silicon and in smaller amounts aluminium, iron,
magnesium, phosphorus, sodium and manganese. These elements are all in
oxidised form, e.g. Na2O, CaO, K2O, but can be reactive or soluble in water to
varying degrees. It is the ash fraction that provides the liming effects of
biochar that is discussed as a potential mechanism of some reported
increases in plant productivity (see Section 3.3). However, for soils that are
salinised or are sensitive to become salinised, that same ash fraction might
pose an increased threat. Surprisingly little work has been found on biochar
ash and under what conditions it may become soluble and contribute to
salinisation.

4.8 Summary
This chapter has described the interactions between biochar and ‘threats to
soil’. For most of these interactions, the body of scientific evidence is currently
insufficient to arrive at a consensus. However, what is clear is that biochar
application to soils will effect soil properties and processes and thereby
interact with threats to soil. Awareness of these interactions, and the
mechanisms behind them, is required to lead to the research necessary for
arriving at understanding mechanisms and effects on threats to soil, as well
as the wider ecosystem.

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5. WIDER ISSUES
5.1 Emissions and atmospheric pollution
The high load of aerosol and pollutant emissions generated by wildfires and
the combustion of fossil fuels explain much of the concern on biochar
production being associated to high levels of particulate matter and
atmospheric pollutants. Nevertheless, the type and composition of such
emissions, including the way these are influenced by pyrolysis conditions and
factors associated to biomass feedstock, are considerably less well
understood (Fernandes and Brooks, 2003).
Particulate matter emitted during pyrolysis is a main focus of human and
environmental health concern based on what is known regarding the inherent
toxicity associated to some types of fine and ultrafine particles, due to their
small size and large surface area (Fernandes and Sicre, 1999). Whereas until
recently, some cases of disease (e.g. respiratory and cardiac) associated to
atmospheric pollution were thought to be caused by some particle types with
dimensions up to 10 µm, recent progress has demonstrated that those
responsible are mainly within the nano-size range. The U.S.A. Environment
Protection Agency (EPA) has responded by putting forward new ambient
standards on Air Quality for particulate matter <2.5 µm (PM2.5). Current
annual mean limits are 40 µg m-3 and 20 µg m-3 for PM10 (<10 µm) and
PM2.5 respectively (EPA, 2007), whereas ambient standards for sub-micron
particles in the environment were not found. Besides the potential health risks
associated to fine and ultrafine particle emissions, their direct and indirect role
in climate change has also granted them wide attention. Further research
involving characterisation of biochar-related particulate emissions during
pyrolysis would be vital for assessing the true contribution of such emissions
to ambient aerosols, as well as identifying processing conditions and
technologies that may help reducing them.
Typically, large amounts of organic and inorganic volatile compounds are
emitted during biomass pyrolysis, particularly at temperatures exceeding
500°C (Greenberg et al., 2005; Gaskin et al., 2008; Chan and Xu, 2009).
Major volatile organic compounds emissions from pyrolysis (30 to 300oC) of
leaf and woody plant tissue (pine, eucalyptus and oak wood, sugarcane and
rice) included acetic acid, furylaldehyde, methyl acetate, pyrazine, terpenes,
2,3 butadione, phenol and methanol, as well as smaller quantities of furan,
acetone, acetaldehyde, acetonitrile and benzaldehyde (Greenberg et al.,
2005). At treatment temperatures between 300 and 600°C, heat- and mass-
transfer rates are high, resulting in a gas-forming pathway dominating the
pyrolysis process, being linked to the production of heavy molecular weight
(tarry) vapours of highly diverse composition (Amonette and Joseph, 2009). At
temperatures around that lower limit, these tars remain trapped within
micropores of the carbonaceous residue but become volatile for higher
temperatures. While the majority of such vapours are commonly recovered
from the gas stream as bio-oil using a condensation tower (Amonette and
Joseph, 2009), a significant proportion is still emitted into the atmosphere,
especially where simple charcoal kilns are used.

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Emissions of PAHs resulting from both natural (e.g. forest fires, volcanic
eruptions) and anthropogenic sources (e.g. burning of fossil fuels) are
recognized as relevant environmental pollutants (Pakdel and Roy, 1991).
Secondary chemical reactions during thermal degradation of organic material
at high temperatures (>700°C), is generally associated to the generation and
emission of heavily condensed and highly carcinogenic and mutagenic PAHs
(Ledesma et al., 2002; Garcia-Perez, 2008). Nevertheless, some evidence
also exists that PAHs can be formed within the temperature range of pyrolysis
(350-600°C). These low-temperature generated PAHs are highly branched in
nature and appear to carry lower toxicological and environmental implications
(Garcia-Perez, 2008). Preliminary results from a recent study have shown that
the amount of biochar-related PAH emissions from traditional feedstocks
remain within environmental compliance (Jones, 2008).
Dioxins (PCDD) and furans (PCDF) are planar chlorinated aromatic
compounds, which are predominantly formed by combustion of organic
material in the presence of chlorine and metals, at temperatures exceeding
1000°C (Lavric et al., 2005; Garcia-Perez, 2008). Wood (accidental fires,
wildfires and wood wastes) is an important air emission source for dioxins
(Lavric et al., 2005). While combustion of firewood and pellets in residential
stoves, as well as paper and plastic wastes, are well know for emitting high
loads of dioxins (Hedman et al., 2006), actual emission factors and
corresponding activity rates remain poorly assessed (Lavric et al., 2005). No
experimental evidence was found confirming dioxin emissions from pyrolysis
of traditional biomass feedstocks used in biochar production.
The emission of atmospheric pollutants during biochar production requires a
full evaluation. This assessment is vital for establishing whether such
emissions may cancel out benefits such as carbon sequestration potential.
Such an evaluation should focus beyond a qualitative and quantitative
characterisation of those pollutants, and should include the pyrolysis
operational conditions and technologies required to reduce their emissions yto
acceptable levels. Evidence in the literature suggests that a certain degree of
control in respect to biochar-related emissions can be achieved through the
use of traditional feedstock materials and lower (<500°C) temperature
pyrolysis. Whereas this aspect looks promising in relation to Air Quality,
current biochar-producing technologies remain largely inefficient. According to
Brown (2009), there is still wide room for improvement in the context of both
energy consumption and atmospheric emissions, particularly when traditional
gasifiers are concerned. At this level, the author identifies specific goals for
optimal biochar production, among which are the use of continuous feed
pyrolisers and an effective recovery of co-products (Brown, 2009). A detailed
analysis on current and future biochar technologies aiming for a more
‘environmentally friendly’ biochar production is also provided.
Collison et al. (2009) in a report to EEDA, reminded that generation and
emission of environmental pollutants as well as the incidence of health and
safety issues associated to biochar production, transport and storage, is
probably of greater concern for small-scale pyrolysis units, particularly in
developing countries. It is often the case, that such smaller units lack the
knowledge and/or financial support, to comply to the environmental standards
(Brown, 2006). A joint effort is necessary to overcome this gap, which

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includes the use of clean pyrolysis technologies (Lehmann et al., 2006) and
the establishment of tight policy and regulations in respect to biochar
production and handling. Furthermore, adequate educating and training, and
perhaps the granting of governmental financial support would allow putting in
place equipment and measures, aiming to minimise environmental and
human exposure to emissions linked to biochar production.

5.2 Occupational health and safety

Biochar production facilities, as well as those associated to transportation and
storage may pose an Occupational Health hazard for the workers involved,
particularly when exposure to biochar dust is concerned (Blackwell et al.,
2009). In addition, health and fire hazards are related directly to the key
physical properties of biochar determining the suitability for a given application
method (Blackwell et al., 2009). However, any discussions and
recommendations in the context of health and safety can only be addressed
generally, given the heterogeneity among biochars. Further research on acute
and chronic exposure to biochar dust, in particular to its nano-sized fraction,
remains scarce and is thus identified as a priority.
‘Nanoparticle’ has been used broadly to refer to those particles within biochar
dust (e.g. fullerenes or fullerene-like structures, crystalline forms of silica,
cristobalite and tridymite), with at least one dimension smaller than 100 nm.
Two major aspects distinguish them from the remaining larger-sized
microparticles: large surface area and high particle number per unit of mass,
which may signify a 1000-fold enhanced reactive surface (Buzea et al., 2007).
Such reactivity and their small size widely explain their hazardous potential.
Several reports have focused on their ability to enter, transit within and
damage living cells and organisms. This capacity is partly consequence of
their small size, enabling easy penetration through physical barriers,
translocation trough the circulatory system of the host, and interaction with
various cellular components (Buzea et al., 2007), including DNA (Zhao et al.,
2005).
Most toxicological and epidemiological studies using fish, mice and
mammalian cell lines (Andrade et al., 2006; Moore et al., 2006; Oberdorster et
al., 2006; Nowack et al., 2007) demonstrate an inflammatory response in the
cell or animal host (Donaldson et al., 2005). In biological systems,
nanoparticles are known to generate disease mainly by mechanisms of
oxidative stress, either by introducing oxidant species into the system or by
acting as carriers for trace metals (Oberdorster, et al., 2004; Sayes et al.,
2005). Those studies have also demonstrated that oxidative stress may result
ultimately in irreversible disruption of basic cellular mechanisms such as
proliferation, metabolism and death. However, extrapolating such effects to
humans remains a challenge, and any outcomes are expected to be
dependent on various factors relating to exposure conditions, residence time
and inherent variability of the host (Buzea et al., 2007).
Exposure to nanoparticles within biochar dust (e.g. carbon-based NP,
crystalline silica) appears to have associated health risks primarily for the
respiratory system (e.g. Borm et al., 2004; Knaapen et al., 2004) and the
gastrointestinal tract (e.g. Hussein et al., 2001). If inhalation of biochar dust
should occur, measures which rapidly enhance airway clearance (e.g.

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mucociliary rinsing with saline solution), and reduce inflammatory and allergic
reactions (e.g. sodium cromoglycate) should be promptly carried out (Buzea
et al., 2007). On the other hand, dermal uptake of combustion-derived
nanoparticles was also found to occur, although this issue remains a
controversial one. It has been suggested that nanoparticle incursion through
the skin may occur at hair follicles (Toll et al., 2004), as well as broken
(Oberdörster et al., 2005) or flexed (Tinkle et al., 2003) skin, depending
mainly on particle size.
Besides unusually high levels (up to 220 g kg-1) of silica, highly toxic
crystalline forms of cristobalite and tridymite have also been found in rice husk
biochars produced at temperatures above 550°C. Blackwell et al. (2009) did
not hesitate in recommending careful handling, transport and storage of rice
husk biochar as well as strict quality control measures for its production.
Regarding those mineral forms, Stowell and Tubb (2003) have recommended
maximum exposure limits of 0.1, 0.05 and 0.05 mg m-3 for crystalline silica,
cristobalite and tridymite respectively. In comparison, those authors have
suggested that current maximum exposure limits for crystalline silica (given as
an example) assigned by the UK (0.3 mg m-3) and the US (10 mg m-3 divided
by the percentage of SiO2) may be too high.
In the context of Occupational Health, reducing biochar dust exposure
requires tight health and safety measures to be put in place. For biochars
containing a large proportion of dust, health risks associated to safe transport
and storage, as well as application, may be reduced using dust control
techniques (Blackwell et al., 2009). For example, covering or wrapping
biochar heaps or spraying the surface with stabilising solutions can minimise
the risk of exposure during transport and storage. In regard to reducing dust
formation during application, especially with concern to uniform topsoil mixing
and top-dressing, water can be used to support on-site spreading (when
spreading is appropriate) (Blackwell et al., 2009).
It has been reported that generation of free-radicals during thermal
(120°C<T<300oC) degradation of lignocellulosic materials, may be
responsible for the propensity of fresh biochars to spontaneously combust
(Amonette and Joseph, 2009), particularly at temperatures <100°C (Bourke et
al., 2007). The free-radicals are primarily produced by thermal action on the
O-functionalities and mineral impurities within the source material. Under
certain conditions, an excessive accumulation of free-radicals at the biochar
surface (Amonette and Joseph, 2009) and within its micropores (Bourke et al.,
2007) might occur. The proportion of free-radicals in biochar is primarily
dependent on the temperature of pyrolysis, and generally decrease with
increasing operation temperatures (Bourke et al., 2007).
There is also evidence that an excessive accumulation of biochar dust in
enclosed spaces may enhance its pyrophoric potential, as recently reported
with coal dust in mines (Giby et al., 2007). To tackle this issue, increasing
biochar density through pelleting may be advisable (Werther et al., 2000). In
addition, the volatile (e.g. aldehydes, alcohols and carboxylic acids) content of
biochar (as influenced by biomass feedstock and operation conditions; Brown
2009) may also constitute a fire hazard during transport, handling and storage
(Werther et al., 2000), and should be taken into account.

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Overall, increasing awareness of biochar flammability means that avoiding
biochar storage with neighbouring residential buildings and goods is
advisable. Nevertheless, successful attempts to reduce the risk of combustion
of rice husk char by adding fire retardants (e.g. boric acid, ferrous sulphate;
Maiti et al., 2006) and inert gases for removal of atmospheric O2 (Naujokas,
1985) have been reported. There is also sound proof of the effective use of
water in assisting cooling of a wide range of carbonaceous materials,
including charcoals (Naujokas, 1985).

5.3 Monitoring biochar in soil

Research methodologies for comparing different biochars produced under
laboratory conditions already have been put in place, based on work involving
charcoal and other BCs. Currently, 13C nuclear magnetic resonance (NMR)
and mid-infrared spectroscopy appear to be reliable methods for providing
compositional characterisation (at the functional group level) of biochar, as
well as differentiation between biochar products. Nevertheless, using such
methods for routine purposes is expensive and time consuming, particularly
when a large number of samples is involved. An efficient, rapid and
economically feasible method for long-term routine assessment of biochar in
soil has not yet been described. Furthermore, at the present, it is perhaps
more important for research to focus on assessing and comparing between
biochar produced under industrial and field conditions.

5.4 Economic Considerations

There is no established business model in the sense of industry-wide
accepted set of standards of production, distribution and use of biochar. In
fact, even the term “biochar industry” would be misplaced. What exists
currently is a multitude of start-up companies and other entities experimenting
with alternative pyrolysis technologies operating at various scales.
Two important considerations with respect to the operation of any biochar
system are: the scale of the biochar operation, and how the feedstock is
sourced (intentional or dedicated). Biochar can be produced in a centralised,
industrial fashion, or can adopt a small-scale, local approach. Regarding
feedstocks, one can distinguish between an open and a closed system. In a
closed system, the pyrolised material essentially consists of agricultural and
forestry residues (byproduct), whereas the open system envisages the
growing of biomass dedicated to pyrolysis as well as off-site waste products
(e.g. sewage sludge). The distinction along these lines is important because
of the different economic implications associated with the respective biochar
systems and it also gives rise to another distinction between private and
social costs and benefits.

5.4.1 Private costs and benefits

The private costs and benefits determine the commercial viability of any
biochar operation and are a combination of biochar’s value as a soil additive,
as a source of carbon credits and as an energy source. Crudely, the cost-
revenue structure of a biochar system could be broken down as follows
(McCarl et al., 2009; Collison et al., 2009).On the revenue side, the following
sources of value should be considered:

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• sale of pyrolysis-derived energy co-products;
• value of biochar as a soil amendment;
• value of biochar as a source of carbon credits.
Potential value to farmers, if any, could arise from increases in crop yield,
although current evidence indicates a relatively small overall effect (see
Section 3.3) and plant production is likely to vary considerably for
combinations of environmental factors and crop types (see Sections 3.3).
Additional economic benefits, in the form of reduced production costs, may
also come about from a reduction in fertilizer application or liming (both very
dependent on biochar quality and quantity as well as frequency of application,
see Section 1.8). Irrigation costs could also potentially be reduced if biochar
application leads to enhanced water retention capacity, which evidence
suggests may be possible at least for sandy soils (see Section 3.1.2).
However, although the intention of biochar is to improve the soil it can also be
envisaged that unforeseen effects on the soil, due to improper management,
would actually lead to an increase in production costs.
For example, when (sub)soil compaction is caused during biochar application
to the soil, subsequent subsoiling operations to alleviate the compaction
would incur a cost. Due to the lack of a functioning biochar industry, it is not
yet clear whether any payments for carbon credits will accrue to the land
owners or the biochar producers. Either way, the economic viability of the
carbon offsetting potential could be limited owing to the potentially high
monitoring and verification costs (Gaunt and Cowie, 2009). Regardless whom
the proceeds from carbon credits accrue to, their value should reflect not only
the carbon sequestration potential of biochar but also the reduced emissions
due to lower fertiliser applications, as well as emissions from the
transportation needs of biomass and biochar. Accounting for these indirect
emissions might add to the costliness of certifying any carbon credits and,
thus, further undermine its profitability.
The cost elements of the equation are the following:
• cost of growing the feedstock (in case of an open system);
• cost of collecting, transporting and storing the feedstock;
• cost of pyrolysis operation (purchase of equipment, maintenance,
depreciation, labour);
• cost of transporting and applying the biochar
Despite the large uncertainties on biochar costs and benefits, the following
factors ought to be taken into account. First, it is clear that the private costs
and benefits of a biochar operation will vary depending on the scale of the
operation. Biochar production at an industrial scale implies significantly higher
costs of transporting the feedstock and the biochar produced from it than
when produced at a small scale. System analysis studies will be of great help
in understanding these issues. Higher transportation needs also lead to higher
GHG emissions, as more fuel is needed for hauling the biomass and the
biochar. The increased emissions need to be accounted for and included in
the carbon offsetting potential of biochar, which would reduce the biochar’s

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value as a source of carbon credits. On the other hand, industrial production
of biochar means that bigger pyrolysis plants could generate economies of
scale, which would bring the average cost of producing biochar down.
Another factor that may influence the commercial appeal and the reliability in
the supply of biochar is the fact that biochar is only one co-product of
pyrolysis, the other ones being syngas and bio-oil. Different types of pyrolysis
(fast vs. slow) will yield different proportions of these products (see Section
1.6), and biochar with varying properties, for a given amount of feedstock.
This means that decisions pertaining to the quantity and quality of produced
biochar will depend on the economic attractiveness of the other two products
and not just on the cost elements of biochar production and the demand for
biochar. For instance, if demand for bio-oil and syngas increases, the
opportunity cost of biochar production will increase, thus shifting production
away from it and rendering it relatively more expensive. Such flexibility in
production is, of course, a welcome trait for pyrolysis operators, but adds an
extra layer of unpredictability that might dampen demand for biochar as a soil
amendment and as a potential source of carbon credits.
As biochar development and adoption are still at an early stage, there is
currently very little quantitative information on these costs and benefits.
McCarl et al. (2009) undertook a cost benefit analysis (CBA) of a pyrolysis
operation in Iowa that uses maize crop residues as feedstock. Assuming a 5 t
ha-1 biochar application and a 5% increase in yields, they conclude that both
fast and slow operations are not profitable at current carbon and energy
prices, with a net present value of about -$44 and -$70 (per tonne of
feedstock) respectively.

Figure 5.1 Effect of transportation distance in biochar systems with bioenergy production using
the example of late stover feedstock on net GHG, net energy and net revenue (adopted from
Roberts et al., 2009)

Roberts et al. (2009) calculate the economic flows associated with the
pyrolysis of three different feedstocks (stover, switchgrass and yard waste).
They find that the economic profitability depends very much on the assumed

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value of sequestered carbon. At $20 t-1 CO2e, only yard waste makes
pyrolysis operation profitable, whereas at a higher assumed price of $80 t-1
CO2e, stover is moderately profitable ($35 t-1 of stover), yard waste
significantly so ($69 t-1 of waste), but switchgrass is still unprofitable. The
point that is made is that despite the revenues from the biochar and energy
products for all feedstocks, the overall profitability is reduced by the cost of
feedstock collection and pyrolysis, even when CO2 is valued at $80 t-1, while
the costs of feedstock and biochar transport and application play a smaller
role. Figure 5.1 illustrates the effect that increased transportation distance has
on net GHG, net energy and net revenue for a pyrolysis operation using
stover as a feedstock.
In a somewhat less sophisticated attempt to estimate costs and benefits,
Collison et al. use a hypothetical case study of biochar application in the East
of England, without, however, taking into account the costs of biochar
production, distribution and application. They estimate an increase in
profitability of the order of £545 ha-1 for potatoes and £143 ha-1 for feed
wheat.
Similarly, Blackwell et al. (2007) estimated the wheat income benefits for
farmers in Western Australia by carrying out a series of trials of applying
varying rates of mallee biochar and fertiliser. The trials produced benefits of
up to $96 ha-1 of additional gross income at wheat prices of $150 ha-1. Again,
no account was taken of the costs of biochar production.
The lesson to be taken from such studies is that at this early stage, any CBA
is an assumption-laden exercise that is prone to significant errors and
revisions as more information becomes available on pyrolysis technologies
and the agronomic effects of biochar.

5.4.2 Social costs and benefits

The social costs and benefits closely follow from the private ones but can be
quite hard to monetize, or even model. Like the private ones, they also
depend on the type of biochar system that is adopted. If an open system is
adopted, the biggest concern is that the drive for larger volumes of biochar
may lead to unsustainable land practices, causing significant areas of land to
be converted into biomass plantations. Such competition for land could
encourage the destruction of tropical forests directly or indirectly, via the
displacement of agricultural production. The latter possibility could also have
negative consequences on the prices and the availability of food crops, much
like in the case of the market for biofuels.
However, these social costs are not inevitable. Tropical deforestation could be
avoided if, for instance, biomass is grown sustainably on land previously
deforested. Moreover, any adverse effects of growing biochar feedstock on
food security and availability could be mitigated by the biochar-induced gains
in crop yields (see Section 3.3). Furthermore, wide, health-related social
benefits can be ascribed to biochar’s potential for land remediation and
decontamination. Of course, the biggest source of social benefits would be
biochar’s climate change mitigation potential.
This section has briefly sketched the economic considerations that ought to
be taken into account when planning for the development of a biochar system.

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For biochar to be successful it must not only deliver on its environmental
promise but it should also be commercially viable.
The profitability of any biochar operation will depend mainly on its potential to
attract revenue as a soil additive and carbon sink and will be affected by the
type of production (open vs. closed, local vs. centralised), which can in turn
result in environmental and economic spillovers. Moreover, the demand for
biochar will be influenced by, and will indeed influence the demand for
biofuels, as a byproduct of pyrolysis, the demand for products such as
manure and compost and the price of carbon in the carbon markets.
Which shape and direction the biochar industry is likely to take is very much
unknown at this stage. However, any outcomes will be greatly influenced by
policy measures on energy, agriculture and climate change. The interplay and
interdependence of such policies call for a holistic, systemic assessment of
the opportunities and pitfalls presented by biochar.

5.5 Is biochar soft geo-engineering?

Geo-engineering is the artificial modification of Earth systems to counteract
the consequences of anthropogenic effects, such as climate change. Large-
scale (industrial) deployment of biochar thus qualifies as a geo-engineering
scheme. Geo-engineering is very controversial and the primitive nature of
geo-engineering schemes has been likened to a planetary version of 19th
century medicine (Lovelock, 2007). Furthermore, panaceas often fail (Ostrom
et al., 2007). However, biochar may be considered a ‘softer’ form of geo-
engineering compared to more intrusive schemes. Especially if used with
certain feedstocks under certain conditions and compared to those geo-
engineering proposals that focus on lowering temperature rather than
reducing GHG emissions or sequestering carbon. Indeed, biochar has been
promoted as a lower-risk strategy compared to other sequestration methods
(Lehmann, 2007). Nevertheless, deploying biochar on a scale with a
mitigative effect entails a large construction of necessary infrastructure and a
very intrusive impact on the way agriculture is performed.
The scalability of biochar is both a potential strength and a potential
weakness. As noted by Woods et al. (2006) ‘one is sometimes left the
impression that the biochar initiative is solely directed towards agribusiness
applications’. However, several trials exist in collaboration with smallholder
farmers, the closest approximation to the original Terra Preta formation. Small
scale biochar systems that lead to a reduction of net GHG emissions have
been suggested to be part of C offset mechanisms and so possibly contribute
to soil C storage in Africa (Whitman and Lehmann, 2009). However, given the
extensive use of biomass burning for energy in Africa, one of the potential
problems will relate to the willingness of farmers to forego an energy source
(biochar) once it has been created, which requires transparent certification
and monitoring schemes if it is to be used in C credit trading schemes.
To what extent are the motives, practices and input materials that led to the
creation of the Terra Preta soils similar or different compared to today’s
application of biochar to soil? A first obvious difference relates to the variety of
inputs used in the formation of Terra Preta, compared to the limited number of

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inputs (e.g. biochar, or mixtures of biochar and manure) currently proposed.
This is an important consideration that determines how far the carbon storage
properties (relative to ‘average’ agricultural soil with organic matter) and
agronomic benefits of Terra Preta can reasonably be extrapolated.
The recalcitrance of biochar components is estimated to be potentially
hundreds or thousands of years (dependent on biochar properties,
environmental conditions, and land use/soil management), or roughly one to
two magnitudes higher than the breakdown of OM in the soil (Sections 3.2.1
and 3.2.5.1). Biochar has been identified as the oldest fraction of SOM,
confirming it recalcitrance to decomposition and mineralisation (Lehman and
Sohi, 2007). The residence time and stability of biochar in Terra Preta soil are
fairly robust, but are the result of extensive smallholder agriculture over tens
to hundreds of years as opposed to intensive agriculture. The direct
translation of these residence times to today’s intensive agricultural systems
with the use of heavy machinery, and the possible accelerated disintegration
and decomposition of biochar particles, with possible effects on biochar
recalcitrance, remains questionable.
Sequestering carbon with biochar seems to have potential in theory. Choices
of feedstocks are critically related to the larger scale impacts and benefits of
biochar. Use of specific organic waste (e.g. papermill waste) may be a
reasonable first approach that circumvents the food vs. fuel debate (cf.
biofuels, van der Velde et al., 2009). Hansen et al. (2008), using illustrative
climate change mitigation scenarios, assumed waste-derived biochar to
provide only a small fraction of the land use related CO2 drawdown, with
reforestation and curtailed deforestation providing a magnitude more
(Kharecha and Hansen, 2009). In line with estimates by Lehman et al. (2006),
Hansen et al. (2008) assumed waste-derived biochar to “be phased in linearly
over the period 2010-2020, by which time it will reach a maximum uptake rate
of 0.16 Gt C yr--1”. This illustrates that waste-derived biochar can be a part of
the mitigation options, although fundamental uncertainties associated with
biochar remain.

5.6 Summary
Biochar can be produced from a wide range of organic feedstocks under
different pyrolysis conditions and at a range of scales. The original feedstock
used, combined with the pyrolysis conditions will affect the exact physical and
chemical properties of the final biochar, and ultimately, the way and the extent
to which soil dependent ecosystem services are affected. Preliminary
evidence appears to suggest that a tight control on the feedstock materials
and pyrolysis conditions (mainly temperature) may be enough in attenuating
much of the current concern relating to the high levels of atmospheric
pollutants (e.g. PAHs, dioxins) and particulate matter that may be emitted
during biochar production, while implications to human health remain mostly
an occupational health issue. Health (e.g. dust exposure) and fire hazards
associated to production, transport, application and storage need to be
considered when determining the suitability of the biochar for a given
application, while tight health and safety measures need to be put in place to
mitigate such risks for the worker, as well as neighbouring residential areas.

118
The profitability of any biochar operation will depend mainly on its potential to
attract revenue as a soil additive and C sink and will be affected by the type of
biomass feedstock and that of production (open vs closed, local vs
centralised), which can, in turn, result in environmental and economic
spillovers. Moreover, the demand for biochar, as a byproduct of pyrolysis, will
be influenced by, and will indeed influence, the demand for biofuels, the
demand for products such as manure and compost and the price of carbon in
the carbon markets. Furthermore, the costs and benefits of a range of biochar
operations and scenarios need to be quantified. Cost-benefit analyses ought
to cast the net wide by accounting not only for commercial factors but also for
social costs and benefits.

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6. KEY FINDINGS
This chapter summarises the main findings of the previous chapters,
synthesises between these and identifies the key research gaps.

6.1 Summary of Key Findings

This report has highlighted that large gaps in knowledge still exist regarding
the effects (including the mechanisms involved) of biochar incorporation into
soils. Considerable further research is required in order to maximise the
possible advantages of such an application, while minimizing any possible
drawbacks. For some potential effects very few or no data are available. For
other effects data exist but they do not cover sufficiently the variation in
relevant soil-environment-climate-management factors. Table 6.1 provides an
overview of the key findings. In view of this, the possibility of qualifying
biochar for carbon offset credits within the UNFCC as part of a post-Kyoto
treaty seems premature at the present stage. Although an inclusion in the
carbon credit systems would certainly boost the nascent biochar industry,
current scientific knowledge of large-scale use of biochar in intensive
agricultural systems has not reached a sufficient level for safe deployment.
Best practices associated with production and application, quality standards,
specifications that clarify land use conflicts and opportunities, monitoring of
utilisation, and details on minimal qualification requirements for certification of
biochar products, require further understanding of the C-sequestration
potential and behaviour of biochar in the environment.

Table 6.1 Overview of key findings (numbers in parentheses refer to relevant sections)

Description Conditions

Empirical evidence of Biochar analogues (pyrogenic BC and charcoal) are found in

substantial quanities in soils of most parts of the world (1.2-1.4)
charcoal in soils exists (long
term)

The principle of improving Anthrosols can be found in many parts of the world, although
normally of very small spatial extent. Contemplation of Anthrosol
soils has been tried
generation at a vast scale requires more comprehensive, detailed
successfully in the past and careful analysis of effects on soils as well as interactions with
other environmental components before implementation (1.2-1.3
and throughout)

Plant production has been Studies have been reported almost exclusively from tropical regions
Positives

with specific environmental conditions, and generally for very limited

found to increase
time periods, i.e. 1-2 yr. Some cases of negative effects on crop
significantly after biochar production have also been reported (3.3).
addition to soils

Liming effect Most biochars have neutral to basic pH and many field experiments
show an increase in soil pH after biochar application when the initial
pH was low. On alkaline soils this may be an undesirable effect.
Sustained liming effects may require regular applications (3.1.4)

High sorption affinity for Biochar application is likely to improve the overall sorption capacity
of soils towards common anthropogenic organic compounds (e.g.
HOC may enhance the
PAHs, pesticides and herbicides), and therefore influence toxicity,
overall sorption capacity of transport and fate of such contaminants. Enhanced sorption
soils towards these trace capacity of a silt loam for diuron and other anionic and cationic
contaminants herbicides has been observed following incorporation of biochar
from crop residues (3.2.2)

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 Microbial habitat and Biochar addition to soil has been shown to increase microbial
biomass and microbial activity, as well as microbial efficieny as a
provision of refugia for
measure of CO2 released per unit microbial biomass C. The degree
microbes whereby they are of the response appears to be dependent on nutrient avaialbility in
protected from grazing soils

Increases in mycorrhizal Possibly due to: a) alteration of soil physico-chemical properties; b)

indirect effects on mycorrhizae through effects on other soil
abundace which is linked to
microbes; c) plant–fungus signalling interference and detoxification
observed increases in plant of allelochemicals on biochar; or d) provision of refugia from fungal
productivity grazers (3.2.6)

Increases in earthworm Earthworms have been shown to prefer some soils amended with
biochar than those soils alone. However, this is not true of all
abundance and activity
biochars, particularly at high application rates (3.2.6)

The use of biochar Charcoal in Terra Preta soils is limited mainly to Amazonia and have
received many diverse additions other than charcoal. Pyrogenic BC
analogues for assessing
is found in soils in many parts of the world but are of limited
effects of modern biochars feedstock types and pyrolysis conditions (Chapter 1)
is very limited

Soil loss by erosion Top-dressing biochar to soil is likely to increase erosion of the
biochar particles both by wind (dust) and water. Many other effects
of biochar in soil on erosion can be theorised, but remain untested
at present (4.1)

Soil compaction during Any application carries a risk of soil compaction when performed
under inappropriate conditions. Careful planning and management
application
could prevent this effect (4.6)

Risk of contamination Contaminants (e.g. PAHs, heavy metals, dioxins) that may be
present in biochar may have detrimental effects on soil properties
Negatives

and functions. The ocurrence of such compounds in biochar is likely

to derive from either contaminated feedstocks or the use of
processing conditions that may favour their production. Evidence
suggests that a tight control over the type of feedstock used and
o
lower pyrolysis temperatures (<500 C) may be sufficient to reduce
the potential risk for soil contamination (3.2.4)

Residue removal Removal of crop residues for use as a feedstock for biochar
production can forego incorporation of the crop residue into the soil,
potentially leading to multiple negative effects on soils (3.2.5.5)

Occupational health and fire Health (e.g. dust exposure) and fire hazards associated to the
production, transport, application and storage of biochar need to be
hazards
considered when determining the suitability for biochar application.
In the context of occupational health, tight health and safety
measures need to be put in place in order to reduce such risks.
Some of these measures have already proved adequate (5.2)
-1
Reduction in earthworm High biochar application rates of >67 t ha (produced from poultry
litter) were shown to have a negative effect on earthworm survival
survival rates (limited
rates, possibly due to increases in pH or salt levels (3.2.6)
number of cases)

Empirical evidence is Biochar analogues do not exist for many feedstocks, or for some
modern pyrolysis conditions. Biochar can be produced with a wide
extremely scarce for many
variety of properties and applied to soils with a wide variety of
modern biochars in soils properties. Some short term (1-2 yr) evidence exists, but only for a
under modern arable small set of biochar, environmental and soil management factors
management and almost no data is available on long term effect (1.2-1.4)
Unknown

C Negativity The carbon storage capacity of biochar is widely hypothesised,

although it is still largely unquantified and depends on many factors
(environmental, economic, social) in all parts of the life cycle of
biochar and at the several scales of operation (1.5.2 and Chapter 5)

Effects on N cycle N2O emissions depend on effects of biochar addition on soil

hydrology (water-filled pore volume) and associated microbial
processes. Mechanisms are poorly understood and thresholds
largely unknown (1.5.2)

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Biochar Loading Capacity BLC is likely to be crop as well as soil dependent leading to potential
incompatibilities between the irreversibility of biochar once applied
(BLC)
to soil and changing crop demands (1.5.1)

Environmental behaviour The extent and implications of the changes that biochar undergoes
in soil remain largely unknown. Although biochar physical-chemical
mobility and fate
properties and stabilization mechanisms may explain biochar long
mean residence times in soil, the relative contribution of each factor
for its short- and long-term loss has been sparsely assessed,
particularly when influenced by soil environmental conditions. Also,
biochar loss and mobility through the soil profile and into the water
resources has been scarcely quantified and transport mechanisms
remain poorly understood (3.2.1)

Distribution and availability Very little experimental evidence is available on the short- and long-
term occurrence and bioavailability of such contaminants in biochar
of contaminants (e.g. heavy
and biochar-enriched soil. Full and careful risk assessment in this
metals, PAHs) within context is urgently required, in order to relate the bioavailability and
biochar toxicity of the contaminant to biochar type and 'safe' application
rates, biomass feedstock and pyrolysis conditions, as well as soil
type and environmental conditions (3.2.4)

Effect on soil organic matter Various relevant processes are acknowledged but the way these are
influenced by combinations of soil-climate-management factors
dynamics
remains largely unknown (Section 3.2.5)

Pore size and connectivity Although pore size distribution in biochar may significantly alter key
soil physical properties and processes (e.g. water retention,
aeration, habitat), experimental evidence on this is scarce and the
underlying mechanisms can only be hypothesised at this stage (2.3
and 3.1.3)

Soil water Adding biochar to soil can have direct and indirect effects on soil
water retention, which can be short or long lived, and which can be
retention/availability
negative or positive depending on soil type. Positive effects are
dependent on high applications of biochar. No conclusive evidence
was found to allow the establishment of an unequivocal relation
between soil water retention and biochar application (3.1.2)

Soil compaction Various processes associated with soil compaction are relevant to
biochar application, some reducing others increasing soil
compaction. Experimental research is lacking. The main risk to soil
compaction could probably be reduced by establishing a guide of
good practice regarding biochar application (3.1.1 and 4.6)

Priming effect Some inconclusive evidence of a possible priming effect exists in

the literature, but the evidence is relatively inconclusive and covers
only the short term and a very restricted sample of biochar and soil
types (3.2.5.4)

Effects on soil megafauna Neither the effects of direct contact with biochar containing soils on
the skin and respiratory systems of soil megafanua are known, nor
the effects or ingestion due to eating other soil organisms, such as
earthworms, which are likely to contain biochar in their guts (3.2.6.3)

Hydrophobicity The mechnanisms of soil water repellency are understood poorly in

general. How biochar might influence hydrophobicity remains largely
untested (3.1.2.1)

Enhanced decomposition of It is unknow how much subsequent agricultural management

practices (planting, ploughing, etc.) in an agricultural soil with
biochar due to agricultural
biochar may influence (accelerate) the disintegration of biochar in
management the soil, thereby potentially reducing its carbon storage potential
(3.2.3)

Soil CEC There is good potential that biochar can improve the CEC of soil.
However, the effectiveness and duration of this effect after addition
to soils remain understood poorly (2.5 and 3.1.4)

Soil Albedo That biochar will lower the albedo of the soil surface is fairly well
established, but if and where this will lead to a substantial soil
warming effect is untested (3.1.3)

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6.1.1 Background and Introduction
As a concept biochar is defined as ‘charcoal (biomass that has been
pyrolysed in a zero or low oxygen environment) for which, owing to its
inherent properties, scientific consensus exists that application to soil at a
specific site is expected to sustainably sequester carbon and concurrently
improve soil functions (under current and future management), while avoiding
short- and long-term detrimental effects to the wider environment as well as
human and animal health'. Inspiration is derived from the anthropogenically
created Terra Preta soils (Hortic Anthrosols) in Amazonia where charred
organic material plus other (organic and mineral) materials appear to have
been added purposefully to soil to increase its agronomic quality. Ancient
Anthrosols have been found in Europe as well, where organic matter (peat,
manure, ‘plaggen’) was added to soil, but where charcoal additions appear to
have been limited or non-existent. Furthermore, charcoal from wildfires
(pyrogenic black carbon - BC) has been found in many soils around the world,
including European soils where pyrogenic BC can make up a large proportion
of total soil organic carbon.
Biochar can be produced from a wide range of organic feedstocks under
different pyrolysis conditions and at a range of scales. Many different
materials have been proposed as biomass feedstocks for biochar. The
suitability of each biomass type for such an application is dependent on a
number of chemical, physical, environmental, as well as economic and
logistical factors. The original feedstock used, combined with the pyrolysis
conditions will determine the properties, both physical and chemical, of the
biochar product. It is these differences in physicochemical properties that
govern the specific interactions which will occur with the endemic soil biota
upon addition of biochar to soil, and hence how soil dependent ecosystem
functions and services are affected. The application strategy used to apply
biochar to soils is an important factor to consider when evaluating the effects
of biochar on soil properties and processes. Furthermore, the biochar loading
capacity of soils has not been fully quantified, or even developed
conceptually.

6.1.2 Physicochemical properties of Biochar

Biochar is comprised of stable carbon compounds created when biomass is
heated to temperatures between 300 to 1000°C under low (preferably zero)
oxygen concentrations. The structural and chemical composition of biochar is
highly heterogeneous, with the exception of pH, which is tipically > 7. Some
properties are pervasive throughout all biochars, including the high C content
and degree of aromaticity, partially explining the high levels of biochar’s
inherent recalcitrance. Neverthless, the exact structural and chemical
composition, including surface chemistry, is dependent on a combination of
the feedstock type and the pyrolysis conditions (mainly temperature) used.
These same parameters are key in determining particle size and pore size
(macro, meso and micropore; distribution in biochar. Biochar's physical and
chemical characteristics may significantly alter key soil physical properties
and processes and are, therefore, important to consider prior to its application
to soil. Furthermore, these will determine the suitability of each biochar for a
given application, as well as define its behaviour, transport and fate in the

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environment. Dissimilarities in properties between different biochar products
emphasises the need for a case-by-case evaluation of each biochar product
prior to its incorporation into soil at a specific site. Further research aiming to
fully evaluate the extent and implications of biochar particle and pore size
distribution on soil processes and functioning is essential, as well as its
influence on biochar mobility and fate.

6.1.3 Effects on soil properties, processes and functions

This section has highlighted the relative paucity of knowledge concerning the
specific mechanisms behind the reported interactions of biochar within the soil
environment. However, while there is still much that is unknown, large steps
have been taken towards increasing our understanding of the effects of
biochar on soil properties and processed. Biochar interacts with the soil
system on a number of levels. Sub-molecular interactions with clay and silt
particles and SOM occur through Van der Waals forces and hydrophobic
interactions. It is the interactions at this scale which will determine the
influence of biochar on soil water repellency and also the interactions with
cations and anions and other organic compounds in soil. These interactions
are very char specific, with the exact properties being influenced by both the
feedstock and the pyrolysis conditions used.
There has been some evidence to suggest that biochar addition to soil may
lead to loss of SOM via a priming effect in the short term. However, there is
only very little research reported in the literature on this subject, and as such it
is a highly pertinent area for further research. The fact that Terra Pretas
contain SOM as well as char fragments seems to demonstrate that the
priming effect either does not exist in all situations or if it does, perhaps it only
lasts a few seasons and it appear not to be sufficient to drive the loss of all
native SOM from the soil. Biochar has the potential to be highly persistent in
the soil environment, as evidenced both by its presence in Terra Pretas, even
after millennia, and also as evidenced by studies discussed in this section.
While biochars are highly heterogeneous across scales, it seems likely that
properties such as recalcitrance and effects on water holding capacity are
likely to persist across a range of biochar types. It also seems probable, that
while difference may occur within biochars on a microscale, biochars
produced from the same feedstocks, under the same pyrolysis conditions are
likely to be broadly similar, with predictable effects upon application to soil.
What remains to be done are controlled experiments with different biochars
added to a range of soils under different environmental conditions and the
precise properties and effects identified. This will lead towards biochars
possibly being engineered for specific soils and climate where specific effects
are required.
After its initial application to soil, biochar can function to stimulate the edaphic
microflora and fauna due to various substrates, such as sugars, which can be
present on the biochar's surface. Once these are metabolised, biochar
functions more as a mineral component of the soil rather than an organic
component, as evidenced by its high levels of recalcitrance meaning that it is
not used as a carbon source for respiration. Rather, the biochar functions as a
highly porous network the edaphic biota can colonise. Due to the large
inherent porosity, biochar particles in soil can provide refugia for

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microorganisms whereby they may often be protected from grazing by other
soil organisms which may be too large to enter the pores. This is likely to be
one of the main mechanisms by which biochar-amended soils are able to
harbour a larger microbial biomass when compared to non-biochar amended
soils. Biochar incorporation into soil is also expected to enhance overall
sorption capacity of soils towards trace anthropogenic organic contaminants
(e.g. PAHs, pesticides, herbicides), in a stronger way, and mechanistically
different, from that of native organic matter. Whereas this behaviour may
greatly contribute to mitigating toxicity and transport of common pollutants in
soil, biochar aging over time may result in leaching and increased
bioavailability of such compounds. On the other hand, while the feasibility for
reducing mobility of trace contaminants in soil might be beneficial, it might
also result in their localised accumulation, although the extent and
implications of this have not been experimentally assessed.
Soil quality may not be necessarily improved by adding biochar to soil. Soil
quality can be considered to be relatively high for supporting plant production
and provision of ecosystem services if it contains carbon in the form of
complex and dynamic substances such as humus and SOM. If crop residues
are used for biochar, the proportion of carbon going into the dynamic SOM
pool is likely to be reduced, with the carbon being returned to the soil in a
relatively passive biochar form. The proportion of residues which are removed
for pyrolysis versus the proportion which is allowed to remain in the soil will
determine the balance between the dynamic SOM and the passive biochar
and so is likely to affect soil quality for providing the desired roles, be it
provision of good use as crop or timber, or functioning as a carbon pool.
Biochar also has the potential to introduce a wide range of hazardous organic
compounds (e.g. heavy metals, PAHs) into the soil system, which can be
present as contaminants in biochar that has been produced either from
contaminated fedstocks or under processing conditions which favour their
production. While a tight control over the feedstock type and processing
conditions used can reduce the potential risk for soil contamination,
experimental evidence of the occurrence and bioavailability and toxicity of
such contaminants in biochar and biochar-enriched soil (over time) remain
scarce. A comprehensive risk assessment of each biochar product prior to its
incoporation into soil, which takes into account the soil type and
environmental conditions, is therefore, paramount.
Increased crop yields are the most commonly reported benefits of adding
biochar to soils. A full search of the scientific literature led to a compilation of
studies used for a meta-analysis of the effects of biochar application to soils
and plant productivity. Meta-analysis techniques (Rosenberg et al., 1997)
were used to quantify the effect of biochar addition to soil on plant productivity
from a range of experiments. Our results showed a small overall, but
statistically significant, positive effect of biochar application to soils on plant
productivity in the majority of cases, covering a range of both soil and crop
types. The greatest positive effects were seen on acidic free-draining soils
with other soil types, specifically Calcarosols showing no significant effect. No
statistically significant negative effects were found. There was also a general
trend for concurrent increases in crop productivity with increases in pH up on
biochar addition to soils. This suggests that one of the main mechanisms

126
behind the reported positive effects of biochar application to soils on plant
productivity may be a liming effect. These results underline the importance of
testing each biochar material under representative conditions (i.e. soil-
environment-climate-management factors).
The degree and possible consequences of the changes biochar undergo in
soil over time remain largely unknown. Biochar loss and mobility through the
soil profile and into water resources has so far been scarcely quantified and
the underlying transport mechanisms are poorly understood. This is further
complicated by the limited amount of long-term studies and the lack of
standardised methods for simulating biochar aging and for long-term
environmental monitoring.

6.1.4 Biochar and soil threats

This chapter has described the interactions between biochar and ‘threats to
soil’. For most of these interactions, the body of scientific evidence is currently
insufficient to arrive at a consensus. However, what is clear is that biochar
application to soils will effect soil properties and processes and thereby
interact with threats to soil. Awareness of these interactions, and the
mechanisms behind them, is required to lead to the research necessary for
arriving at understanding mechanisms and effects on threats to soil, as well
as the wider ecosystem.

6.1.5 Wider issues

Biochar can be produced from a wide range of organic feedstocks under
different pyrolysis conditions and at a range of scales. The original feedstock
used, combined with the pyrolysis conditions will affect the exact physical and
chemical properties of the final biochar, and ultimately, the way and the extent
to which soil dependent ecosystem services are affected. Preliminary
evidence appears to suggest that a tight control on the feedstock materials
and pyrolysis conditions (mainly temperature) may be enough in attenuating
much of the current concern relating to the high levels of atmospheric
pollutants (e.g. PAHs, dioxins) and particulate matter that may be emitted
during biochar production, while implications to human health remain mostly
an occupational health issue. Health (e.g. dust exposure) and fire hazards
associated to production, transport, application and storage need to be
considered when determining the suitability of the biochar for a given
application, while tight health and safety measures need to be put in place to
mitigate such risks for the worker, as well as neighbouring residential areas.
The profitability of any biochar operation will depend mainly on its potential to
attract revenue as a soil additive and C sink and will be affected by the type of
biomass feedstock and that of production (open vs closed, local vs
centralised), which can, in turn, result in environmental and economic
spillovers. Moreover, the demand for biochar, as a byproduct of pyrolysis, will
be influenced by, and will indeed influence, the demand for biofuels, the
demand for products such as manure and compost and the price of carbon in
the carbon markets. Furthermore, the costs and benefits of a range of biochar
operations and scenarios need to be quantified. Cost-benefit analyses ought
to cast the net wide by accounting not only for commercial factors but also for
social costs and benefits.

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6.2 Synthesis
The aim of this report was to review the state-of-the-art regarding the
interactions between biochar application to soils and effects on soil properties,
processes and functions. Adding biochar to soil is not an alternative to
reducing the emissions of greenhouse gasses. Minimising future climate
change requires immediate action to lower greenhouse gas emissions and
harness alternative forms of energy (IPCC, 2007).

6.2.1 Irreversibility
The irreversibility of biochar application to soils has implications for its
development. Once biochar has been applied to soils, it is virtually impossible
to remove. This irreversibility does not have to be a deterrent from considering
biochar. Rather, the awareness of its irreversibility should lead to a careful
case-by-case assessment of its impacts, underpinned by a comprehensive
body of scientific evidence gathered under representative soil-environment-
climate-management conditions. Meta-analyses, an example of which on the
relationship between biochar and crop productivity is presented in this report,
can provide a valuable method for both signalling gaps in knowledge as well
as providing a quantitative review of published experimental results. The
results of meta-analyses can then be used to feed back to directing funding
for more research where needed, and/or to inform specific policy
development. Objectivity of systematic reviews on biochar is of paramount
importance. In the medical sciences this has been resolved by the founding of
an independent organisation (the Cochrane Collaboration), which provides
regularly updated systematic reviews on specific healthcare issues using a
global network of volunteers and a central database/library. A similar
approach, although at a different scale, could be envisaged to ensure that the
most robust and up to date research informs policy concerning biochar.
Alternatively, this task could be performed by recognised, independent
scientific institutions that do not (even partially) depend on conflicting funding,
and that have the necessary expertise.

6.2.2 Quality assessment

The evidence reviewed in this report has highlighted potential negative as well
as positive effects on soils and, importantly, a very large degree of unknown
effects (see Table 6.1; and Section 6.3). Some of the potential negative
effects can be ‘stopped at the gate’, i.e. by not allowing specific feedstocks
that have been proven to be inappropriate, and by regulating pyrolysis
conditions to avoid undesirable biochar properties (a compulsory biochar
quality assessment and monitoring approach could prove effective). Other
potential negative effects on soils, or the wider ecosystem, need to be
regulated on the application side, i.e. at the field scale, taking into account the
soil properties and processes as well as threats to soil functions. Similarly,
biochar properties can be ‘engineered’ (to an extent), through controlled use
of feedstocks and pyrolysis conditions, to provide necessary benefits to soil
functions and reduce threats when applied to fields that have specific soil-
environmental-climatic-management conditions. However, the current state-
of-the-art regarding the effects of biochar on soils has a substantial lack of

128
information on relevant factors (see Section 6.3). Results from research into
the relative importance of these factors, and the associated environmental
and soil management conditions, needs to drive further extension and
development of a biochar quality assessment protocol.

6.2.3 Scale and life cycle

Relevant factors for producing biochar with specific properties are feedstock
characteristics and pyrolysis conditions, thereby affecting the scale and
method of operation. The optimal scale of operation, from a soil improvement
and climate adaptation perspective, will differ for different locations, as the
availability of feedstocks and the occurrence of soil-environment-climate-
management conditions changes along with land use. The optimal scale of
operation, from a climate mitigation perspective, is, intuitively, the smallest
scale. However, full life cycle assessment studies to evidence this have not
been found. It is possible that at a larger scale of operation, if not production
then at least application, a more complementary situation exists with larger
concomitant reductions in CO2 equivalent emissions by the ability to forego or
reduce certain operations. For example, a farm on a fertile floodplain, with
good water availability, may produce biochar from feedstocks on the farm with
good water and nutrient retention properties. If this is applied to soils on the
same farm, it may allow a reduction of a single fertiliser pass. However, if the
biochar is sold (or traded) to the farm next door, which may be on soils with
low water and nutrient retention, then there may be a reduction of two fertiliser
passes and a substantial reduction in irrigation, for example. It is possible,
therefore, that the CO2 equivalents saved on the farm next door are more
than the CO2 equivalent emissions produced during transport from one farm
to the other. This is of course just one hypothetical example of how off-site
biochar distribution does not necessarily decrease the carbon negativity of the
technology. One critical factor affecting this is the way long-lived specific
beneficial effects of specific biochars will be under specific conditions.
Experimental studies of sustained effects, e.g. nutrient and water retention, of
different biochars in different soil-environment-climate-management
combinations are needed to feed into life cycle assessment studies. It is
possible that the optimum scale of operation, in terms of global warming
mitigation, will be different in different parts of Europe and the world.

6.2.4 Mitigation/adaptation
Besides global warming mitigation, biochar can also be viewed from the
perspective of adaptation to climate change. In the future, climate change
looks likely to increase rainfall intensity, if not annual totals, for example
thereby increasing soil loss by water erosion, although there is much
uncertainty about the spatio-temporal structure of this change as well as the
socio-economic and agronomic changes that may accompany them.
Independent from changes in climate, the production function of soil will
become increasingly more important, in view of the projected increase in
global human population and consequent demands for food. More than 99%
of food supplies (calories) for human consumption come from the land,
whereas less than 1% comes from oceans and other aquatic ecosystems
(FAO, 2003).

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A common way of thinking about adapting food production to climate change
is by genetically engineering crops to survive and produce under adverse and
variable environmental conditions. This may well work, if risks to the
environment are minimised and public opinion favourable. However, other soil
functions are likely to still be impaired and threats exacerbated, such as
increased loss of soil by erosion. Improving the properties of soil will increase
the adaptive capacity of our agri-environmental systems. The ClimSoil report
(Schils et al., 2008) reviews in detail the interrelation between climate change
and soils. One of their conclusions is that land use and soil management are
important tools that affect, and can increase, SOC stocks. In this way, the
soils will be able to function better, even under changing climatic conditions.
In arable fields, SOM content is maintained in a dynamic equilibrium. Arable
soil is disturbed too much for it to maintain greater contents of SOM than a
specific upper limit, which is controlled by mainly clay contents and the soil
wetness regime. Biochar, because of its recalcitrance, and possibly because
of its organo-mineral interaction and accessibility, provides a means of
potentially increasing the relevant functions of soils beyond that which can be
achieved by OM alone in arable systems.
Biochar application to soils, therefore, may play both a global warming
mitigation and a climate change adaptation role. For both, more research is
needed before conclusive answers can be given with a high degree of
scientific certainty, particularly when considering specific soil-environment-
climate-management conditions and interactions. However, it may be the
case that in certain situations the biochar system does not mitigate global
warming, i.e. is C neutral or positive, but that the enhanced soil functions from
biochar application may still warrant contemplation of its use.
As far as the current scientific evidence allows us to conclude, biochar is not a
‘silver bullet’ or panacea for the whole host of issues ranging from food
production and soil fertility to mitigating (or more correctly ‘abating’) global
warming and climate change for which it is often posited. The critical
knowledge gaps are manifold, mainly because the charcoal-rich historic soils,
as well as most experimental sites, have been studied mostly in tropical
environments, added to the large range of biochar properties that can be
produced from the feedstocks currently available subjected to different
pyrolysis conditions. Biochar analogues, such as pyrogenic BC, are found in
varying, and sometimes substantial amounts in soils all over the world. As
well as causing some difficulty with predicting possible impacts of biochar
addition to soil, the large variety in biochar properties that can be produced
actually provides an opportunity to ‘engineer’ biochar for specific soil-
environment-climate-management conditions, thereby potentially increasing
soil functioning and decreasing threats to soil (and/or adapting to climate
change). What is needed is a much better understanding of the mechanisms
concerning biochar in soils and the wider environment. Although the research
effort that would be required is substantial, the necessary methods are
available.

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6.3 Knowledge gaps
Table 6.1 lists ‘unknown’ effects of biochar on soil properties, processes and
functions. For ‘known’ positive or negative effects, Table 6.1 also discusses
(briefly but with reference to more elaborate discussions in the report) the soil-
environment-climate-management conditions for which the effects are valid
and where they are not (known). From the viewpoint of biochar effects on soil
functions and soil threats, a number of key issues emerge that are discussed
in the subsections below. Biochar research should aim to reach a sufficient
level of scientific knowledge to underpin future biochar policy decisions. This
review indicates that a large number of questions related to biochar
application to soils remain unanswered. The multitude of gaps in current
knowledge associated with biochar properties, the long-term effects of biochar
application on soil functions and threats, and its behaviour and fate in different
soil types (e.g. disintegration, mobility, recalcitrance, interaction with SOM), as
well as sensitivity to management practices, require more scientific research.

6.3.1 Safety
While the widespread interest in biochar applications to soils continues to rise,
issues remain to be addressed concerning the potential for soil contamination
and atmospheric pollution associated to its production and handling, with
potentially severe health, environmental and socio-economic implications. The
irreversibility of biochar incorporation into soil emphasises the urgent need for
a full and comprehensive characterisation of each biochar type in regard to
potential contaminants (mainly heavy metals and PAHs), as influenced by
biomass feedstock and pyrolysis conditions. Very little focus has been paid to
the long-term distribution of such contaminants in biochar-enriched soils and
bioavailability to the micro- and macro-biota. In this context, risk assessment
procedures for these compounds need to be re-evaluated on a case-by-case
basis, based on bioavailable concentrations (rather than initial concentrations
in biochar) and accounting for the influence of NOM on their desorption from
biochar over time. This would allow understanding the true implications of
their presence in biochar on human, animal and ecosystem health over a wide
range of soil conditions, while enabling relation of toxicity to biochar type and
safe application rates, as well as feedstock characteristics and pyrolysis
conditions. Similarly, the emission of atmospheric pollutants during biochar
production requires careful qualitative and quantitative analysis. It will provide
a sound basis for the development and/or optimisation of feedstock and
pyrolysis operational conditions (as well as technologies) required to tackle
these pollutants.

6.3.2 Soil organic matter dynamics

Biochar can function as a carbon sink in soils under certain conditions.
However, the reported long residence times of biochar have not been
confirmed for today’s intensive agricultural systems in temperature regions.
Disintegration of biochar is likely to be stimulated by intensive agricultural
practices (tilling, plouging, harrowing) and use of heavy machinery, thereby
potentially reducing residence times. Work is required to better elucidate the
biochar loading capacity of different soils, for different climatic conditions in
order to maximise the amount of biochar which can be stored in soils without
impacting negatively on soil functions. In addition to crop yields, research

131
should also focus on threshold amounts of biochar that can be added to soils
without adverse consequences to soil physical properties, such as priming by
increasing the pH or dedcreasing water-filled pore space, hydrophobic effects,
or soil chemical properties, e.g. adding a high ash content (with salts) biochar
to a soil already at risk of salinisation, or other ecosystem components, e.g.
particulate or dissolved organic C reaching ground/surface waters. Therefore,
the biochar loading capacity should vary according to environmental
conditions as well as biochar ‘quality’, specific to the environmental conditions
of the site (soil, geomorphology, hydrology, vegetation).

6.3.3 Soil biology

Owing to the vital role that the soil biota plays in regulating numerous
ecosystem services and soil functions, it is vital that a full understanding of the
effects of biochar addition to soil is reached before policy is written. Due to the
very high levels of heterogeneity found in soils, with regard to soil physical,
chemical and biological properties, extensive testing is needed before
scientifically sound predictions can be made regarding the effects of biochar
addition to soils on the native edaphic communities under a range of climatic
conditions. Much of the data currently reported in the literature shows a slight,
but significant positive effect on the soil biota, with increased microbial
biomass and respiration efficiency per unit carbon, with associated increases
in above ground biomass production reported in the majority of cases. There
is currently a major gap in our understanding of the influence of biochar
addition to soils on carbon fluxes. This is vital to increase our understanding
of interactions between the soil biota and biochar as it will help to unravel the
mechanisms behind any possible priming effect, as well as nutrient transfer
and interactions with contaminants introduced with biochar. A very suitable
method for probing this interaction would be the use of Stable Isotope Probing
(SIP), which can be used with other molecular techniques to trace the flow of
carbon from particular sources through the soil system. Pyrolysing biomass
labeled with a stable isotope and measuring its emission from the soil will
allow accurate measures of its recalcitrance over time. Conducting controlled
atmosphere experiments with stable isotope-labelled CO2 will enable
assessing the observed increased microbial respiration and investigation of
whether this increase is due to a more efficient use of plant provided
substrates (in case the label is detected in soil respiration), or if a priming
effect has occurred leading to increased metabolisation of the SOM (in case
the label is not detected).

6.3.4 Behaviour, mobility and fate

Physical and chemical weathering of biochar over time has implications for its
solubilisation, leaching, translocation through the soil profile and into water
systems, as well as interactions with other soil components (including
contaminants). Up to now, biochar loss and environmental mobility have been
quantified scarcely and such processes remain poorly understood. In addition,
the contribution of soil management practices and the effects of increasingly
warmer climates, together with potential greater erosivity as potential key
mechanisms controlling biochar fate in soil, have also been assessed
insufficiently up to now.

132
An effective evaluation of the long-term stability and mobility of biochar,
including the way these are influenced by factors relating to biochar
physicochemical characteristics, pyrolysis conditions and environmental
factors, is paramount to understanding the contribution that biochar can make
to improving soil processes and functioning, and as a tool for sequestering
carbon. Such knowledge should derive from long-term studies involving a
wide range of soil conditions and climatic factors, while using standardised
methods for simulating biochar aging and for long-term environmental
monitoring.

6.3.5 Agronomic effects

Biochar has shown merit in improving the agronomic and environmental value
of agricultural soils in certain pilot studies under limited environmental
conditions, but a scientific consensus on the agronomic and environmental
benefits of biochar has not been reached yet. It remains difficult to generalise
these studies due to the variable nature of feedstocks, their local availability,
the variability in resulting biochar and the inherent biophysical characteristics
of the sites it has been applied to, as well as the variability of agronomic
practices it could be exposed to. Furthermore, there is a lack of (long-term)
studies on the effects of biochar application in temperate regions. Direct and
indirect effects of biochar on soil hydrology (e.g. water availability to plants)
need to be studied experimentally for representative conditions in the field and
in the laboratory (soil water retention – pF - curves) before modelling
exercises can begin. Ultimately, in those conditions where biochar application
is beneficial to agriculture and environment, it should be considered as part of
a soil conservation package aimed at increasing the resilience of the agro-
environmental system combined with the sequestration of carbon. The key is
to identify the agri-soil management strategy that is best suited at a specific
site. Other carbon sequestration and conservation methods, such as no-till,
mulching, cover crops, complex crop rotations, mixed farming systems and
agroforestry, or a combination of these, need to be considered. In this context
the interaction of biochar application with other methods warrants further
investigation.

133
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European Commission

EUR 24099 - EN – Joint Research Centre – Institute for Environment and Sustainability
Title: Biochar Application to Soils - A Critical Scientific Review of Effects on Soil Properties,
Processes and Functions
Author(s): F. Verheijen, S. Jeffery, A.C. Bastos, M. van der Velde, I. Diafas
Luxembourg: Office for Official Publications of the European Communities
2009 – 151 pp. – 21.0 x 29.7 cm
EUR – Scientific and Technical Research series – ISSN 1018-5593
ISBN 978-92-79-14293
DOI 10.2788/472

Abstract
Biochar application to soils is being considered as a means to sequester carbon (C) while
concurrently improving soil functions. The main focus of this report is providing a critical
scientific review of the current state of knowledge regarding the effects of biochar application
to soils on soil properties and functions. Wider issues, including atmospheric emissions and
occupational health and safety associated to biochar production and handling, are put into
context. The aim of this review is to provide a sound scientific basis for policy development,
to identify gaps in current knowledge, and to recommend further research relating to biochar
application to soils. See Table 1 for an overview of the key findings from this report. Biochar
research is in its relative infancy and as such substantially more data are required before
robust predictions can be made regarding the effects of biochar application to soils, across a
range of soil, climatic and land management factors.

Definition
In this report, biochar is defined as: “charcoal (biomass that has been pyrolysed in a zero or
low oxygen environment) for which, owing to its inherent properties, scientific consensus
exists that application to soil at a specific site is expected to sustainably sequester carbon
and concurrently improve soil functions (under current and future management), while
avoiding short- and long-term detrimental effects to the wider environment as well as human
and animal health." Biochar as a material is defined as: "charcoal for application to soils". It
should be noted that the term 'biochar' is generally associated with other co-produced end
products of pyrolysis such as 'syngas'. However, these are not usually applied to soil and as
such are only discussed in brief in the report.

Biochar properties
Biochar is an organic material produced via the pyrolysis of C-based feedstocks (biomass)
and is best described as a ‘soil conditioner’. Despite many different materials having been
proposed as biomass feedstock for biochar (including wood, crop residues and manures), the
suitability of each feedstock for such an application is dependent on a number of chemical,
physical, environmental, as well as economic and logistical factors. Evidence suggests that
components of the carbon in biochar are highly recalcitrant in soils, with reported residence
times for wood biochar being in the range of 100s to 1,000s of years, i.e. approximately 10-
1,000 times longer than residence times of most soil organic matter. Therefore, biochar
addition to soil can provide a potential sink for C. It is important to note, however, that there is
a paucity of data concerning biochar produced from feedstocks other than wood, but the
information that is available is discussed in the report. Owing to the current interest in climate
change mitigation, and the irreversibility of biochar application to soil, an effective evaluation
of biochar stability in the environment and its effects on soil processes and functioning is
paramount. The current state of knowledge concerning these factors is discussed throughout
this report.

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Pyrolysis conditions and feedstock characteristics largely control the physico-chemical
properties (e.g. composition, particle and pore size distribution) of the resulting biochar,
which in turn, determine the suitability for a given application, as well as define its behaviour,
transport and fate in the environment. Reported biochar properties are highly heterogeneous,
both within individual biochar particles but mainly between biochar originating from different
feedstocks and/or produced under different pyrolysis conditions. For example, biochar
properties have been reported with cation exchange capacities (CECs) from negligible to
approximately 40 cmolc g-1, and C:N ratios from 7 to 500, while the pH is normally neutral to
basic . While this heterogeneity leads to difficulties in identifying the underlying mechanisms
behind reported effects in the scientific literature, it also provides a possible opportunity to
engineer biochar with properties that are best suited to a particular site (depending on soil
type, hydrology, climate, land use, soil contaminants, etc.).

Effects on soils
Biochar characteristics (e.g. particle and pore size distribution, surface chemistry, relative
proportion of readily available components), as well as physical and chemical stabilisation
mechanisms of biochar in soils, determine the effects of biochar on soil functions. However,
the relative contribution of each of these factors has been assessed poorly, particularly under
the influence of different climatic and soil conditions, as well as soil management and land
use. Reported biochar loss from soils may be explained to a certain degree by abiotic and
biological degradation and translocation within the soil profile and into water systems.
Nevertheless, such mechanisms have been quantified scarcely and remain poorly
understood, partly due to the limited amount of long-term studies, and partly due to the lack
of standardised methods for simulating biochar aging and long-term environmental
monitoring. A sound understanding of the contribution that biochar can make as a tool to
improve soil properties, processes and functioning, or at least avoiding negative effects,
largely relies on knowing the extent and full implications of the biochar interactions and
changes over time within the soil system.

Extrapolation of reported results must be done with caution, especially when considering the
relatively small number of studies reported in the primary literature, combined with the small
range of climatic, crop and soil types investigated when compared to possible instigation of
biochar application to soils on a national or European scale. To try and bridge the gap
between small scale, controlled experiments and large scale implementation of biochar
application to a range of soil types across a range of different climates (although chiefly
tropical), a statistical meta-analysis was undertaken. A full search of the scientific literature
led to a compilation of studies used for a meta-analysis of the effects of biochar application to
soils and plant productivity. Results showed a small overall, but statistically significant,
positive effect of biochar application to soils on plant productivity in the majority of cases. The
greatest positive effects were seen on acidic free-draining soils with other soil types,
specifically calcarosols showing no significant effect (either positive or negative). There was
also a general trend for concurrent increases in crop productivity with increases in pH up on
biochar addition to soils. This suggests that one of the main mechanisms behind the reported
positive effects of biochar application to soils on plant productivity may be a liming effect.
However, further research is needed to confirm this hypothesis. There is currently a lack of
data concerning the effects of biochar application to soils on other soil functions. This means
that although these are qualitatively and comprehensively discussed in this report, a robust
meta-analysis on such effects is as of yet not possible. Table 1 provides an overview of the
key findings - positive, negative, and unknown - regarding the (potential) effects on soil,
including relevant conditions.

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Preliminary, but inconclusive, evidence has also been reported concerning a possible priming
effect whereby accelerated decomposition of soil organic matter occurs upon biochar addition
to soil. This has the potential to both harm crop productivity in the long term due to loss of soil
organic matter, as well as releasing more CO2 into the atmosphere as increased quantities of
soil organic matter is respired from the soil. This is an area which requires urgent further
research.

Biochar incorporation into soil is expected to enhance overall sorption capacity of soils
towards anthropogenic organic contaminants (e.g. PAHs, PCBs, pesticides and herbicides),
in a mechanistically different (and stronger) way than amorphous organic matter. Whereas
this behaviour may greatly mitigate toxicity and transport of common pollutants in soils
through reducing their bioavailability, it might also result in their localised accumulation,
although the extent and implications of this have not been assessed experimentally. The
potential of biochar to be a source of soil contamination needs to be evaluated on a case-by-
case basis, not only with concern to the biochar product itself, but also to soil type and
environmental conditions.

Implications
As highlighted above, before policy can be developed in detail, there is an urgent need for
further experimental research in with regard to long-term effects of biochar application on soil
functions, as well as on the behaviour and fate in different soil types (e.g. disintegration,
mobility, recalcitrance), and under different management practices. The use of representative
pilot areas, in different soil ecoregions, involving biochars produced from a representative
range of feedstocks is vital. Potential research methodologies are discussed in the report.
Future research should also include biochars from non-lignin-based feedstocks (such as crop
residues, manures, sewage and green waste) and focus on their properties and
environmental behaviour and fate as influenced by soil conditions. It must be stressed that
published research is almost exclusively focused on (sub)tropical regions, and that the
available data often only relate to the first or second year following biochar application.

Preliminary evidence suggests that a tight control on the feedstock materials and pyrolysis
conditions might substantially reduce the emission levels of atmospheric pollutants (e.g.
PAHs, dioxins) and particulate matter associated to biochar production. While implications to
human health remain mostly an occupational hazard, robust qualitative and quantitative
assessment of such emissions from pyrolysis of traditional biomass feedstock is lacking.

Biochar potentially affects many different soil functions and ecosystem services, and interacts
with most of the ‘threats to soil’ outlined by the Soil Thematic Strategy (COM (2006) 231). It is
because of the wide range of implications from biochar application to soils, combined with the
irreversibility of its application that more interdisciplinary research needs to be undertaken
before policy is implemented. Policy should first be designed with the aim to invest in
fundamental scientific research in biochar application to soil. Once positive effects on soil
have been established robustly for certain biochars at a specific site (set of environmental
conditions), a tiered approach can be imagined where these combinations of biochar and
specific site conditions are considered for implementation first. A second tier would then
consist of other biochars (from different feedstock and/or pyrolysis conditions) for which more
research is required before site-specific application is considered.

From a climate change mitigation perspective, biochar needs to be considered in parallel with
other mitigation strategies and cannot be seen as an alternative to reducing emissions of
greenhouse gases. From a soil conservation perspective, biochar may be part of a wider
practical package of established strategies and, if so, needs to be considered in combination

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with other techniques.

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