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Greenhouse gas production in mixtures of soil with composted and
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non-composted biochars is governed by char-associated organic
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compounds
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Nils Borchard1,3*, Kurt Spokas2, Katharina Prost1and Jan Siemens1
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1
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University of Bonn, Nussallee 13, 53115 Bonn, Germany
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Institute of Crop Science and Resource Conservation, Soil Science and Soil Ecology,
United States Department of Agriculture-Agricultural Research Service, Soil and Water
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Management Research Unit,1991 Upper Buford Circle, St. Paul, MN 55108 USA
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3
Agrosphere (IBG-3), Jülich Research Centre, 52425 Jülich, Germany
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*Corresponding author,
[email protected], Tel: ++49-2461-616583, Fax: ++49-
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2461-612518
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Abstract
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Biochar application to soil has the potential to increase soil productivity while reducing
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anthropogenic greenhouse gas (GHG) emissions to the atmosphere. However, techniques for
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conditioning this material for maximizing its beneficial effects as a soil amendment still
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require elucidation. We examined changes of organic matter associated with two biochars
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after 175 d composting and resulting effects on GHG emissions during 150 d incubation.
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Composting decreased the amount of organic compounds that could be thermally released
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from the biochars and also affected their molecular nature. These thermally-desorbable
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organic compounds from initial biochars likely stimulated the oxidation of methane and
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inhibited the production of nitrous oxide in soil-biochar mixtures. However, these reductions
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of GHG emissions disappeared together with thermally-desorbable organic compounds
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following the composting of chars. Instead, addition of composted gasification coke and
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charcoal stimulated the formation of methane and increased nitrous oxide emissions by 45 to
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56%. Nitrous oxide emissions equaled 20% of the total amount of N added with composted
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biochars, suggesting that organic compounds and nitrogen sorbed by the chars during
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composting fueled GHG production. The transient nature of the suppression of CH4 and N2O
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production challenges the long-term GHG mitigation potential of biochar in soil.
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Keywords: Biochar, Compost, GHG, thermally-desorbable organic compounds
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1 Introduction Agricultural activities release considerable amounts of greenhouse gases (GHGs) like
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carbon dioxide (CO2), nitrous oxide (N2O), and methane (CH4) to the atmosphere. The N2O
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and CH4 emission from this sector represents up to 50% of the total global anthropogenic N2O
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and CH4 emissions (Cole et al., 1997; IPCC, 2007). Thus, management options that mitigate
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these GHG emissions are attracting research interest (IPCC, 2007; Smith et al., 2008).
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Producing biochar with an output of energy and its subsequent use as soil amendment may at
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least partly offset GHG emissions from agriculture (Laird, 2008). Part of this compensation is
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related to the sequestration of photosynthetically fixed carbon for long periods of time in soil
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and the reduction of CO2, N2O, and CH4 emissions from soil (Libra et al., 2011; Woolf et al.,
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2010). Furthermore, biochar applications can potentially increase soil fertility as well as
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corresponding crop yields (Biederman and Harpole, 2013; Jeffery et al., 2011).
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The term biochar is commonly used for the pyrogenic residue produced by chemical
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and thermal alteration of biomass for soil application to improve soil quality by affecting its
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physical and chemical properties while at the same time sequestering carbon to mitigate
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climate change (Joseph et al., 2010; Montanarella and Lugato, 2013; Woolf et al., 2010). The
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potential of biochars to sequester carbon is related to their ability to resist degradation in soil
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(e.g. Harvey et al., 2012). This resistance against biogeochemical degradation depends on its
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variable contents of stabile and more labile carbon (Hammes et al., 2008; Keiluweit et al.,
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2010; Zimmerman, 2010), with the stabile pool being made up of highly condensed aromatic
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carbon and the labile pool consisting of organic compounds sorbed or otherwise associated
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with this aromatic “backbone” (Harvey et al., 2012; Zimmerman, 2010). Since biochars
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interact with soil constituents and solutes (Ameloot et al., 2013; Cayuela et al., 2014; Fang et
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al., 2013), their application to soils can be expected to affect biogeochemical processes and
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hence GHG emissions from soils. However, the changes of CO2 emissions from soils after application of biochar are
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variable; with positive or negative effects, and no impact in CO2 emissions being observed in
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short-term laboratory incubations (Cross and Sohi, 2011; Kammann et al., 2012; Zimmerman
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et al., 2011). Furthermore, the composition and the content of soil organic matter in the
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receiving soils affects the decomposition of the labile organic matter fraction of the biochars,
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and hence the soil’s CO2 production (Cross and Sohi, 2011; Spokas and Reicosky, 2009).
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Given the greater global warming potentials of CH4 and N2O compared to CO2 (IPCC, 2013),
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researchers have paid more attention to biochar effects on CH4 and N2O emissions from
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amended soils. However, the mechanistic knowledge about CH4 emissions following biochar
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applications to terrestrial soils is still scarce (Libra et al., 2011), while N2O emissions were
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more thoroughly investigated (Cayuela et al., 2013; Cayuela et al., 2014; Harter et al., 2013).
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The response of the soil microbial community to biochar incorporation has also been variable
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with regard to CH4, with both stimulation and suppression of production as possible reaction
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(Libra et al., 2011; Yu et al., 2013). We think that these observed differences are related to the
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variation in the amount and composition of organic compounds associated with the different
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biochars, since some hydrocarbons can stimulate methane oxidation, being a competitive food
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source for the methanotrophs (Hazeu and Bruyn, 1980), or a potential chemical inhibitor of
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enzymatic activities (Hubley et al., 1975) affecting CH4.
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With regard to N2O, it is known that additions of ash-containing (alkaline) biochar to
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an acidic or neutral soil have the potential to increase soil pH, thereby reducing N2O
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emissions (Cayuela et al., 2014; Libra et al., 2011). Furthermore, biochar may sorb NH4+,
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soluble organic nitrogen and to a lesser extend also NO3- (Borchard et al., 2012; Cayuela et
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al., 2014; Prost et al., 2013), which may reduce N2O formation by limiting the availability of
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nitrogen for further nitrification, denitrification, and nitrifier denitrification (Cayuela et al.,
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2014; Libra et al., 2011). Moreover, biochars are highly porous (Joseph et al., 2010; Keiluweit
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et al., 2010) and can therefore provide suitable habitats for N2O-transforming
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microorganisms, create aerobic and/or anaerobic microsites, change soil moisture holding
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capacity, and increase soil aeration (Lehmann et al., 2011; Rogovska et al., 2011; van Zwieten
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et al., 2010). However, regardless of the interaction between soil moisture and biochar effects
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on N2O emissions (Cayuela et al., 2014; Yanai et al., 2008), recent research has suggested that
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soil aeration effects are not critical to N2O suppression mechanisms (Case et al., 2012;
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Cayuela et al., 2013). Furthermore, biochars may initially contain sorbed organic compounds
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that may inhibit microbially mediated processes (Bedmutha et al., 2011; Cayuela et al., 2014;
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Spokas et al., 2011). Recent studies of the abundance of genes involved in the N cycle
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indicate that biochar additions to soils or manure induce shifts in the microbial community
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and activity that can be linked to reductions in N2O emissions (i.e. an increasing abundance of
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nosZ genes, decreasing abundance of nirK genes; Harter et al., 2013; Wang et al., 2013). All
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of these processes can act jointly or in combination resulting in a reduction of N2O production
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(Cayuela et al., 2014; Clough et al., 2013; Yanai et al., 2008).
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The fertilizer value of biochars may be enhanced by composting them with organic
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substrates prior to their addition to soil (Prost et al., 2013; Schulz et al., 2013). During the
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composting process, biochar surfaces are modified due to biotic and abiotic oxidation and
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sorption of compost-derived organic compounds (Borchard et al., 2012; Hua et al., 2009).
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This enrichment of oxygen-containing functional groups on the biochar surface may increase
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NH4+ as well as NH3 retention (Hua et al., 2009). However, it is unknown whether the
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nitrogen retained during composting affects N2O formation in soils after application of
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composted biochars. Moreover, composting is known to decompose low molecular weight
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organic pollutants that may be present in the composting substrate (Semple et al., 2001).
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Composting of biochars could therefore reduce the concentrations of potentially toxic
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compounds and enhance the microbially mediated formation of N2O (Spokas et al., 2011;
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Taghizadeh-Toosi et al., 2011).
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Overall, composting elevates the concentrations of soluble organic N, inorganic N,
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and organic C associated with the biochars, because soluble N and C compounds are released
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from the composting substrate and subsequently sorbed by the biochars. These compounds
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could affect N2O and CO2 production, as well as CH4 formation and oxidation after the
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application of theses composted biochars to soil. Composted biochars should increase N2O
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formation, as they contain smaller amounts of associated toxic or inhibitory organic
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compounds and larger contents of sorbed compost-derived nitrogen and organic carbon. To
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test this hypothesis, we assessed concentrations and composition of thermally-desorbable
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organic compounds of two hardwood-derived biochars prior to and after composting with
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farmyard manure, and quantified GHG production of incubated soil, biochars, and biochar-
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soil mixtures.
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2 Material and Methods
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2.1 Biochars
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We used two biochars for our experiments, produced either by slow pyrolysis
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(commercially available charcoal, proFagus GmbH, Bodenfelde, Germany) or gasification
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(gasification coke, Mothermik GmbH, Pfalzfeld, Germany). Slow pyrolysis charcoal was
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chosen as it has already been used as a reference material to study biochar effects on GHG
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formation in soils before (e.g. Jones et al., 2011; Spokas, 2013). However, we also decided to
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utilize gasification coke, as it is a by-product of the growing bio-fuel industry that will likely
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be available in large quantities for soil application in the future (e.g. Meyer et al., 2011).
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However, the knowledge of gasification coke and its effects on GHG formation after 6
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application to soil is still limited. Detailed information regarding the production process and
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properties of the chars are provided by Borchard et al. (2012) and Prost el al. (2013). Briefly,
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both biochars were produced from hardwood (Fagus and Quercus species). For the
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production of charcoal the wood was carbonized over a period of 13 to 18 h (depending on its
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initial moisture content) at temperatures of up to 550°C. Unlike the production of charcoal the
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wood passed during gasification a downdraft gasifier and was heated to a maximum
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temperature of 1100°C. For the composting experiment the biochars were crushed with a jaw
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crusher (Retsch GmbH, Haan, Germany, Type BB2), sieved to the fraction between 0.125
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mm and 2 mm and filled into litter bags (polyester fabric, size:15 x 15 cm, mesh size: 0.125
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mm), each filled with 50 g biochar.
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2.2 Composting experiment
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We conducted the composting experiment at the experimental farm “Wiesengut” in
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Hennef, Germany (50°47’ N; 7°16’ E), operated by the Institute of Organic Agriculture
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(University of Bonn). The composting setup is described in Borchard et al. (2012) and Prost
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el al. (2013). In brief, we used as composting substrate a mixture of farmyard manure from a
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cow-calf system mixed with wheat straw. We filled the mixtures into twelve separate
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composting containers (volume: 1 m³, material: high-density polyethylene, producer: Mauser
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SM). Each treatment (composting substrate + gasification coke biochar, composting substrate
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+ charcoal, and composting substrate without biochar) was repeated four times and stored
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indoors in a complete randomized design.
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The litter bags were placed separately into the centre of the container (10 litter bags
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per layer in 25, 35, 45, 55, and 65 cm height above ground surface; resulting in 2.5 kg of
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bagged biochar added per container). The composting substrate was remixed with a manure
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spreader every 14 d within the first 84 d and every 28 d up to the end of the experiment (at the
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175th d).The compost temperature of the composting substrate was logged every 30 min (DS 7
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1820; Hygrosens Instruments GmbH, Löffingen, Germany) and the total mass of each
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container was determined before and after turning operations. On day 2, 10, 28, 56, 84, 133
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and 175 of the experiment three litter bags and ~200 g of the compost substrate were removed
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from each container for analysis. After sampling the biochar litter bags as well as the
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farmyard manure and compost were dried at 40°C to a constant weight (2 to 5 d).
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2.3 Chemical and physical characterization of biochars
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Carbon-, H-, and N-contents were measured with a CHN analyzer (EURO EA,
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HEKAtech GmbH, Wegberg, Germany); O-contents were calculated by difference (Borchard
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et al., 2012). Water-extractable organic carbon (WEOC), total soluble nitrogen (TSN) and
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nitrate-nitrogen (NO3-N) were extracted with 0.01 M CaCl2 and measured as described in
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Prost et al. (2013). We assessed the water holding capacity by immersion of the biochars in
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water for 30 min, drainage by gravity for 2 hours, then weighing and subtracting the biochar
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dry weight and expressing this moisture as percent of the biochar dry weight (modified after
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Öhlinger, 1996).
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Thermally-desorbable organic compounds were measured using the headspace-
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thermal desorption method described in Spokas et al. (2011). In brief, 0.5 g of each biochar,
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farmyard manure or compost (from day 2, 10, 28, 56, 84, 133, and 175 of the composting
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process) were placed in a 10 mL headspace vial and sealed with Teflon-lined septa (Grace,
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Deerfield, IL USA). The vial was heated to 75oC for 10 min and the corresponding headspace
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gas was routed through a 10-port diaphragm valve (DV22-2110; Valco Instruments Co., Inc.;
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Houston, TX) to allow the introduction of two simultaneous gas samples from a single
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headspace venting to two different analytical gas chromatography columns. The first was a
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capillary column (RTX-624; 30 m x 0.25 µm; 1.5 mL min-1 He flow rate) housed in a
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temperature controlled GC oven (40 to 275oC at 10oC min-1) and connected to a mass
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spectrometer (MS; Perkin Elmer Clarus T600). The second was a packed column (Porapak Q,
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6.4 mm × 1.8 m; Restek Corp.; Bellefonte, PA) that was connected directly to a thermal
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conductivity detector (TCD), which was in series with a flame ionization detector. Quantified
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compounds were identified by comparing their retention times with liquid injections of
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certified analytical standards into the headspace vial (502.2 MegaMix®, Ketones Mix, 524.2
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Rev. 4.1, Restek Corporation, Bellfaunte, PA), certified gas standards (Minneapolis Oxygen,
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Minneapolis, MN), or diluted neat chemicals (Restek Corporation, Bellfaunte, PA) under
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identical operating conditions of both GC columns and by confirming their mass spectra with
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the on-line NIST library (TurboChrom; NIST library). An 80% match criteria was used for
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NIST library matches. Quantification was accomplished through an external calibration
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method, comparing the response of injected liquid analytical standards to the amount
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observed from the thermal desorption.
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The aromatic backbone of the biochars was analysed using benzene polycarboxylic
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acids (BPCA) as geochemical markers as described in Brodowski et al.(2005), the carbon
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yields from BPCAs with 3 to 6 carboxyl groups were not multiplied with a conversion factor
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(Schneider et al., 2010).
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2.4 Incubation experiment with soil-biochar mixtures
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The impact of the two initial biochars and the composted biochars on GHG and O2
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production or consumption in soil-biochar mixtures was assessed in laboratory incubation
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experiments resembling the approach of Spokas (2013). The agricultural soil used in these
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experiments was collected from the University of Minnesota's Research and Outreach Station
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in Rosemount, MN USA (44°45' N, 93°04' W) and was frequently used to study C cycling
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and effects of biochar on GHG formation in fertile soils (Clay et al., 2007; Spokas, 2013;
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Venterea et al., 2005). The soil is a Waukegan silt loam (Hapludoll in the USDA
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classification and Chernozem in the FAO classification), containing 22% sand, 55% silt, and 9
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23% clay. It is characterized by a pH (1:1 H2O, w/v) of 6.5, 4.4% total organic carbon, and a
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laboratory determined moisture content at field capacity (-33 kPa) of 14.8±0.4% (w/w) for the
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sieved and loosely repacked soil that was incubated. The addition of 10% (w/w) biochar had
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no significant effect on this moisture content at a suction of -33 kPa (15.0±0.7 %). The site
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from which the soil was collected was farmed in a conventionally tilled (moldboard plow)
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corn (Zea mays L.) and soybean (Glycine max (L.) Merr.) rotation. The soil was sampled
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following corn harvest in the fall. Surface soil (0-5 cm) was collected, sieved to < 2 mm, air
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dried, and homogenized for the incubation study, which started three month after soil
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sampling.
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Soil and soil-biochar mixtures were analyzed for major soil nutrients (Ca, Mg, K, and
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P) by Mehlich-3 extraction (Mehlich, 1984). Organic matter of the soil and biochar mixtures
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was estimated through loss on ignition at 500 °C for 2 h(Nelson and Sommers, 1996). In
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addition, the amount of ammonium and nitrate were assessed with 2M KCl extractions (1:20
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w/v; Paul et al., 2004). The KCl extracts were analyzed for NH4-N and NO3-N using an
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automated flow-through injection analyzer (Lachat, Loveland, CO). The potential CEC of the
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soil and soil+biochar mixtures were assessed by the ammonia acetate method buffered at pH 7
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(Sumner and Miller, 1996), which has also been used to analyze other biochar amended soil
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mixtures (Liang et al., 2006).
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Triplicate incubations were established for each initial and composted biochar with the
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following combinations of biochar and soil: 0.5 g biochar, 5 g agricultural soil, and 0.5 g
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biochar + 5 g agricultural soil. Although amending soil with 10% biochar is not realistic for
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agricultural practice, such high amendments are useful for studying basic processes of biochar
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effects on GHG formation in soils (Jones et al., 2011; Spokas, 2013). Corresponding to the
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approach of Spokas and Reicosky (2009), the above incubations were carried out at laboratory
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temperature (22 ± 1.7oC) and at field capacity (-33kPa) of the laboratory soil to establish soil
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moisture as found in winter/spring or following heavy rain events (Harter et al., 2013) to
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investigate biochar effects on N2O and CH4 production at potentially oxygen limited
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conditions. Soil moisture contents equal to field capacity were established by adding 0.74 ml
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of de-ionized water to soil and soil-biochar mixtures. Biochar controls without soil
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(composted and non-composted biochars) also received 0.74 mL of deionized water to
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saturate them to >90% of the water holding capacity of the initial biochars (Table 1). Recent
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research suggests that the interaction between aeration and biochar effects on N2O production
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is small at (water filled pore space >70%; Case et al., 2012; Cayuela et al., 2013). The 125
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mL serum vials (Wheaton Glass, Millville, NJ USA) were sterilized before additions of either
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soil, biochar or soil-biochar mixtures. After adding the respective materials and water to each
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vial it was sealed with a red butyl rubber septa and an aluminium crimp seal (Grace,
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Deerfield, IL USA). Soil and biochar were manually mixed in the serum bottle prior to
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moisture additions. Incubations of biochar (composted and non-composted) without soil were
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conducted to assess the production/consumption of GHG solely from the biochar. Gases from
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incubations were monitored every 3 days in the initial phase of the incubation and every 10
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days in the end and analysed on a gas chromatographic system to quantify average gas
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production rates over a 150 d incubation period (Spokas and Bogner, 2011). However, if the
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O2 level dropped below 15% during the incubation, the incubation was terminated and the
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rates of production were calculated up to this point to maintain comparison of aerobic
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conditions across all incubations.
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The production rate of each gas from the incubations was estimated by the following formula
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(assuming 25oC and 1 atm):
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GHG production rate (g d -1 ) =
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where the slope is the change in gas concentration in the headspace per day (derived from
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fitting a linear equation to all the periodic headspace gas concentrations versus time), MW is
slope (ppmv d -1 ) (MW )(χ ) 120 mL -1 , Vmolar 1000 mL L
(Eq. 1)
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the molecular weight of the gas of interest, and χ is the ratio of the molar mass of carbon or
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nitrogen to molecular weight of the gas (i.e., 12/44 for CO2, 28/44 for nitrogen in N2O; 12/16
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for CH4), Vmolar is the molar volume of a gas (24.465 L mol-1), and the last term is the
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conversion of volume units and accounting for the headspace volume of the serum bottle (120
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mL).
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2.5 Statistical evaluation
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To assess statistically significant impacts of composting on GHG emissions (CO2 and
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N2O production and CH4 oxidation activities) data were analysed using an ANOVA
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procedure for independent samples (MINITAB software, Minitab, Inc., State College, PA). If
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significant differences existed among the factors, as indicated by the F-ratio, the Tukey's
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Honest Significant Difference (HSD) test was performed to determine which pair-wise
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interactions were significantly different at the P
