Contents

1. Introduction ….......................................................….1

2. The potential impact of conservation farming…........3

3. Agroforestry systems ……….........................……....6

4. Biochar as soil amendment ……….........….………...9

5. Regulations ………...........…................................…11

6. Conclusions ……………..............................………13

1 Introduction

Climate change and global warming have worldwide consequences. The most prominent factor driving these phenomena is the increased atmospheric concentrations of greenhouse gases (GHGs) (WMO 2007a). The main GHGs after water vapor are, in descending order, carbon dioxide (CO2), methane (CH4), and nitrous oxide (N2O). The current atmospheric concentrations of CO2, CH4, and N2O are 38%, 158%, and 19% higher than those during the pre-industrial era, reaching 387 ppm (vs. 280 ppm), 1,803 ppb (vs. 700 ppb), and 323 ppb (vs. 270 ppb), respectively (WMO 2010). During the twentieth century, average global surface temperature increased by 0.6 ± 0.12°C and is projected to increase by 1.5–5.8°C by the end of the twenty-first century (IPCC 2007).

Emissions of CO2 are attributed to the increased use of fossil fuels, as well as to the enhanced clearing and burning of forests (Fearnside 2000), and the expanding of agriculture. With the intensification of agriculture, world soils have become a large source of atmospheric CO2, CH4, and N2O (Lal 2002) (Fig. 1). Most agricultural soils have lost from 30% to 75% of their original soil organic C (SOC) pool, namely, 30–40 Mg C ha−1. Consequently, the SOC pool in agricultural soils is much lower than their potential capacity (Lal et al. 2007). Depletion of the SOC pool in arable lands is attributed to lower input of biomass, as well as to higher outputs through oxidation, leaching of dissolved organic carbon (C), and soil erosion (Lal 2004) (Fig. 2). Global CO2 emissions since 1850 are estimated at 78 ± 12 Pg from soils compared with 270 ± 30 Pg from fossil fuel combustion (Lal 2004). Depletion of the SOC pool also degrades soil quality (Stavi and Lal 2011),and decreases crop yields (Lal et al. 2007). N2O is emitted to the atmosphere from a range of natural and anthropogenic sources, including combustion of fossil fuels, biomass burning, and a range of industrial processes (WMO 2007b). Agricultural practices, mainly fertilizer use, also increase N2O emission from the soil. The rate of N2O efflux from soil varies with time, and higher rates are associated with periods of high soil moisture, high temperature, and low soil NO3–N concentration (Perdomo et al. 2009). CH4 is produced naturally from wetlands and through anthropogenic activity, mainly cultivation of paddy rice (Oryza sativa L.), and ruminant husbandry (WMO 2007b; Hobbs et al. 2008). Globally, agricultural lands account for about 25% of the CO2, 50% of the CH4, and 70% of the N2O emissions (Hutchinson et al. 2007).

Fig. 1
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Agriculture intensification and emissions of greenhouse gases from soils. SOC soil organic carbon. Copied with permission from Lal (2002)

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Fig. 2
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Soil organic carbon dynamics. DOC dissolved organic carbon. Copied with permission from Lal (2004)

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Increasing atmospheric concentrations of GHGs have led to a positive radiative forcing of the climate, accelerating warming of the earth surface (IPCC 1995). The radiative forcing has increased by ∼23% from 1990 to 2006. The CO2, CH4 and N2O are responsible for ∼60%, 20%, and 6%, respectively, of the total radiative forcing of the earth’s surface (WMO 2007b). The global warming potential (GWP) of CH4 and N2O emissions is, respectively, 21 and 310 times that of CO2 (Forster et al. 2007). If the CO2 emission rate had remained at levels of the mid-1990s, it would have led to a nearly constant rate of increase in atmospheric concentrations for at least two centuries, reaching about 500 ppm by the end of the twenty-first century. Stabilizing concentrations of GHGs could be achieved by drastically curtailing global anthropogenic emissions. For stabilization of CO2 concentration at 450 or 650 ppm, accumulated anthropogenic emissions over the period 1991–2100 should not exceed 630 or 1,030 Pg C, respectively. Stabilizing atmospheric concentrations of CH4 and N2O at levels of the mid-1990s would require decreased anthropogenic emissions by ∼10% and 50%, respectively (IPCC 1995).

The agronomic and environmental advantages of several conservation agricultural practices including reduced and no tillage, crop residue management, cover cropping, perennial pastures, and aerobic rice are widely recognized. However, most practices have a limited potential of C sequestration over the long run. Furthermore, most of these practices are not applicable under diverse physical and biotic conditions. Thus, this article is focused on two promising practices—agroforestry systems and soil application of biochar. Intensive research and development and wide adoption of these two practices could boost their C sequestration capacity and simultaneously, improve a range of ecosystem services, and advance global food security. Therefore, the objectives of this review were to: (1) demonstrate the climatic-related impacts of several conservation agricultural practices, (2) highlight the mitigation potential of climate change by the practices of agroforestry and biochar, and (3) emphasize the need for well-developed regulations, aimed at a widespread adoption of these practices by the agricultural sector.

2 The potential impact of conservation farming

Compared to intensive agricultural practices, conservation farming can reduce emissions of GHGs. Improved management of agricultural lands may simultaneously decrease fuel inputs during the production processes (Hobbs et al. 2008), increase SOC stocks, and maintain ecosystem services (Stavi and Lal 2010) (Table 1). A range of conservation agricultural methods have been often aggregated under guides for “Best Management Practices” aimed at sustaining soil resources (e.g., Ontario Ministry of Agriculture Food & Rural Affairs’ website: http://www.omafra.gov.on.ca/english/environment/field/fieldcrop.htm), controlling pollution of off-site water sources (e.g., New Hampshire Department of Agriculture Markets and Food 2002) or maintaining air quality (e.g., Governor’s Agricultural Best Practices Committee 2008).

Table 1 Some examples of the effects of conservation agricultural practices on SOC stock, emissions of GHGs from soil, and additional ecosystem services
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Among the important aspects of conservation farming is the decreased disturbance to the structure of the uppermost soil layers. This is achieved through the practicing of reduced tillage or no-till (NT) method of seedbed preparation coupled with crop residue management. The retention of crop residues on the soil surface decreases raindrop impact, lessens aggregate slaking and dispersion (Stavi and Lal 2010), and protects the soil from water and wind erosion (Govaerts et al. 2007). The retention of crop residues also increases infiltration and decreases evaporation loss, favoring crop yields where water availability limits production. The decreased disturbance of the soil profile aids in maintaining its structure, encouraging proliferation and activity of earthworms and other soil’s fauna. These effects promote wildlife biodiversity and support agro-ecosystem health (Huggins and Reganold 2008). At the same time, the reduced disturbance of the soil structure decreases emission rates of CO2, N2O, and CH4 (Ussiri and Lal 2009; Ussiri et al. 2009). The resultant increase in SOC concentration further stimulates soil’s structure formation and stability (Govaerts et al. 2007). In economic terms, the number of passes over a field needed to establish and harvest a crop under NT decreases dramatically, requiring 50–80% less fuel and 30–50% less labor (Huggins and Reganold 2008).

However, only about 7% of the globe’s croplands are managed under NT. Of these, about 85% are in North and South America. At the same time, adoption rates of NT have remained low in Europe, Africa, and Asia. The introduction of conservation agriculture has been especially difficult in developing countries, where crop residues are often used as animal feed, fuel (Huggins and Reganold 2008), or as construction material. Among many subsistence farmers in Africa, immediate problems such as extreme poverty, food insecurity, and poor agricultural productivity are the most common competing needs for crop residues. As a result, mulching materials are often in critically low supply, increasing vulnerability of soils to erosion and degradation. In addition to these, the retention of crop residues is not always feasible, for example, where mulch is consumed by termites during a very short period after harvest (Giller et al. 2009). Furthermore, the conversion to NT from conventional farming may create a number of agronomic, environmental, and economic challenges. For example, different pest species may emerge, exacerbating infestation of soil-borne fungal diseases that tillage previously kept checked. In addition, the conversion increases the need for chemical weed control, aggravating the risks of contamination of soil and water resources. Furthermore, it leads to high costs for purchasing of the required apparatuses, which may be prohibitive specifically for small-scale farms (Huggins and Reganold 2008). Moreover, these conservation practices are not applicable under all circumstances. For instance, intensive tillage may be inevitable in cases of soil hardening following continuous NT under dry climatic conditions (Cockroft and Olsson 2000). In addition, in cases of poor drainage, moist conditions, and temperature regimes of temperate to cool, the elimination of crop residues and the practice of intensive tillage may enhance soil drying and warming, favoring crop production (Vetsch et al. 2007). Therefore, in many long-term NT systems, the implementation of occasional tillage may be required to control spread of pathogens, as well as of weeds (Lal et al. 2007). However, occasional tillage may eliminate some of the advantages achieved by long-term NT farming. For example, in a study conducted in the Midwest region of the USA, Stavi et al. (2011) reported reduced physical quality and hydraulic conductivity, and decreased SOC concentration in the tilled layer following occasional tillage in a long-term NT agro-ecosystem.

Another common conservation practice is cover cropping, which is usually practiced during the “off-season.” Cover cropping protects the ground surface from erosional processes, increases soil porosity and structure formation, recycles residual fertilizers from the soil profile, and adds organic C to the rhizosphere (Reicosky and Forcella 1998). In cases of leguminous cover crops, they can supply considerable stocks of nitrogen (N) to the soil, which are available for uptake by the subsequent main crops (Unger and Vigil 1998). Similar to residue management and conservation tillage, however, cover cropping may also not be feasible for all geographic zones and climatic conditions. For example, under relatively cool temperatures, cover crops may retard emergence and development of the main crops during the spring (Dabney et al. 2001). Under dry climatic conditions, cover cropping may deplete soil moisture reserves and decrease amounts of available water for the main crops (Unger and Vigil 1998).

An additional environmentally friendly practice comprises perennial pastures that are grown as an integral component of the multi-year cropping sequence and include a range of native and introduced grasses, forbs, and legumes for fodder or forage production. The pastures are efficient in reclaiming degraded agricultural lands while providing a range of ecosystem services, such as the increase in species diversity (Sanderson et al. 2004), the support of complex food webs, and the boost in soil health. The deep and dense root systems of pastures support the soil structure formation and C sequestration to a much larger extent than do those of annuals (Ernst and Siri-Prieto 2009; Kell 2011). In a study in Kansas, in the Midwest USA, Culman et al. (2010) studied the impact of perennial grasslands on quality and fertility of the soil up to a 1-m depth. They compared characteristics of soil under native prairie meadows that have never been tilled or fertilized and have been annually harvested for hay for more than 75 years, and of soil under annual cropping systems planted in wheat (Triticum aestivum L.), sorghum (Sorghum bicolor L.), and soybean (Glycine max L.), and chemically fertilized for the past several decades. Culman and colleagues reported that the perennial grasses improved the physical, chemical, and biological quality of the soil. Of special importance were the observed more developed root biomass, greater microbial biomass, more complex food web, higher SOC stock, and increased fertility under the perennial grasses than under the annual cropping systems. In the same region, DuPont et al. (2010) investigated the impact of conversion of a never fertilized, periodically burnt, and annually harvested native tallgrass prairie to a NT annual cropland consisting of soybean, sorghum, and wheat. The conversion was conducted by using herbicides and without tillage. However, DuPont and colleagues reported that the soil under annual crops faced a considerable decrease in readily oxidizable C and in microbial biomass and a substantial modification in soil biota.

In functional terms, a somewhat similar case includes perennial grain crops, such as intermediate wheatgrass (Elytrigia elongata), Aleppo millet grass (Sorghum halepense), and Maximilian sunflower (Helianthus maximiliani), which develop large and dense root systems (Cox et al. 2006). Compared with the annual grain crops, the perennials are characterized as low-input crops, with a small environmental footprint, and with a high C sequestration capacity (Glover et al. 2007; Kell 2011). These crops are therefore capable of maintaining ecosystem services, particularly on marginal lands or where resources are limited (Glover et al. 2010). However, the perennial grain crops’ culture is still in initial stages of development and faces several challenges. Additional research is needed in breeding and husbandry, specifically aimed at increasing crop yields and commercial viability (Cox et al. 2006; Glover et al. 2007, 2010). Shrubby and viney leguminous crops represent an intermediate case of semi-perennials that are vigorous producers of N-enriched vegetation that live longer than the typical annuals (Snapp et al. 2010). In Malawi, Snapp et al. (2010) examined the potential of the grain legume semi-perennials pigeon pea (Cajanus cajan L.), mucuna (Mucuna pruriens L.), and tephrosia (Tephrosia vogelii) in increasing corn (Zea mays) production through intercropping (in space) and rotation (on the temporal axis). Overall, they reported that intercropping had no consistent effect on productivity of the agro-ecosystems. However, in the rotational systems, corn grain yields were similar on a 2-year basis when planted continuously and when rotated with the semi-perennials. At the same time, the nutritional value of the rotational systems was much higher due to the high protein yield of the legumes. In addition, both fertilizer efficiency and stability of yield were greater in the rotational systems than in continuous non-fertilized or fertilized corn.

In regard to rice farming, conversion of paddies to aerobic production systems may considerably decrease CH4 emissions and diminish denitrification losses of N. However, at the same time, such a conversion would considerably increase CO2 emissions from the soil (Smith et al. 2003; Kögel-Knabner et al. 2010). Furthermore, the conversion to aerobic conditions may exacerbate infestation by weeds and pests, augmenting the need of using herbicides and pesticides.

Advances in development of information technologies and related data processing procedures enable the utilization of precision farming systems, aimed at decreasing the farms’ operational costs while reducing the environmental footprint of agriculture (Pierce and Nowak 1999). These advances are especially relevant to nutrient management and can considerably decrease fertilizer input, thereby reducing emissions of N2O (Smith et al. 2008), diminishing eutrophication of off-site water sources, and increasing economic returns (Hatfield 2000). In addition, precision farming is also useful for weed and pathogen control, reducing the economic and environmental costs associated with the utilization of herbicides and pesticides (Pierce and Nowak 1999). The lower energy input on fuels and chemicals also decreases emissions of CO2 at the global level. However, wide adoption of precision agriculture in many developing countries is constrained, as the required information technologies and associated facilities are relatively expensive and not accessible for subsistence farmers.

Overall, it seems that the offset potential of climate change by the above discussed conservation practices is site dependent and vary greatly according to the prevalent conditions. While a specific practice may sequester C and decrease emissions of GHGs under certain conditions, it may increase oxidation of SOC and promote emissions of GHGs under other conditions. Moreover, some of the conservation practices are not applicable under wide ranges of physical and biotic conditions. In addition, to some extent, a sort of tradeoff occurs for these practices between C sequestration and adverse agronomic and environmental impacts. For example, conservation tillage systems that require larger efforts in chemical weed control, which in turn increase the risks of contamination of soil and water resources. Similarly, crop residue management may increase infestation of pests, leading to a greater need in applying pesticides. Likewise, application of manure or compost, despite augmenting the SOC stock, also increases the emissions of GHGs from soil and may result in the pollution and eutrophication of off-site water sources. The highly promising perennial crops, which have been proved to sequester large amounts of C and to accomplish a range of ecosystem services, are more efficient in producing animal feed (hay) than food for humans (grains). Therefore, in order to fulfill the increased demand for food production and at the same time to alleviate the increased pressures on the environment, more focus should be given to additional, specific practices that are highly compatible with these challenges. In the next sections, two such practices—agroforestry systems and the utilization of biochar as soil amendment—are discussed.

3 Agroforestry systems

Agroforestry, an important conservation farming option, encompasses several practices such as alley cropping, multi-storey cropping, and silvopastoral systems, where trees or shrubs are intercropped with grain crops, vegetables, or forages (Kandji et al. 2006) (Fig. 3). Additional agroforestry practices include windbreaks and riparian forests, aimed at fulfilling specific objectives, such as wind erosion control or off-site water quality preservation, respectively (Schoeneberger 2009). Agroforestry systems provide a range of products, including food for human consumption, fodder for livestock, timber for building, wood for fuel, and pollen for honey bees (Nair et al. 2010). At the same time, these systems sustain many ecosystem services such as increasing species diversity, enhancing wildlife habitats, fostering natural food webs, fostering water infiltration, improving soil and ecosystem health, augmenting long-term C sequestration, and decreasing emissions of CO2 (USDA-NRCS 2002; Garrity et al. 2010; Nair et al. 2010). Specifically, these systems are considered as an efficient means in reclaiming degraded lands (Kandji et al. 2006; Soto-Pinto et al. 2010). A wide range of agroforestry systems are presented in Table 2. Despite their being of highly heterogeneous natures, all of these systems reveal a high capacity of sequestering C, along with the improvement of several ecosystem services.

Fig. 3
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Agroforestry system composed of olive trees for oil, intercropped with alfalfa (Medicago sativa L.) for hay. Alfalfa is considered as a perennial forage legume that could last up to several years. At the same time, the forage fertilizes the soil with N and protects the surface from erosional processes. Picture was taken in Yizrael Valley, northern Israel

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Table 2 Impact of agroforestry systems on C stocks and additional ecosystem services
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Successful and well-managed integration of trees on farms may optimize the use of water and nutrients (Sileshi et al. 2011), enabling diversified and sustainable crop production (Kandji et al. 2006). In a study of the nutrients dynamics of an alley cropping system in subtropical China, it was reported that Hog plum (Choerospondias axillaris) trees and a peanut (Archis hypogaea) crop may compete for N fertilizer in the surface soil. However, it was also reported that the trees used leached N in the deep soil not accessed by the shallow-rooted peanuts. The alley cropping system increased N use efficiency compared with a mono-peanut cropping system. These effects increased with time and were more evident for systems with larger trees. These results suggest that the introduction of trees can reduce N loss, but management practices should aim at preventing competition for N with the main crop in the shallow soil depth (Zhang et al. 2008).

Leguminous trees (fertilizer trees) in agroforestry systems can facilitate biological N fixation, augmenting N availability for the main crops, and thereby reducing the requirements for fertilizer inputs (Garrity et al. 2010; Nair et al. 2010). In a study in the sub-Saharan Africa, where rainfall is erratic and highly variable, Sileshi et al. (2011) compared rain use efficiency and crop yield in rain-fed corn when grown as a mono-culture and when intercropped with the N fixation tree Leucaena (Leucaena leucocephala) under two different types of agroforestry systems. Treatments included fully fertilized corn alone, non-fertilized (control) corn alone, and corn intercropped with Leucaena and amended with half of the recommended N fertilizer rate. Sileshi and colleagues reported that despite some discrepancies, rain use efficiency was higher for corn intercropped with Leucaena. In addition, the intercropping increased stability of the corn yield over the years. Overall, these results were true compared with corn alone, both with and without fertilizer. Sileshi and colleagues attributed this to the fact that during extremely wet seasons, mineral fertilizers are subject to leaching, resulting in decreased crop yields. At the same time, the considerably increased SOC stock under leguminous trees enhances nutrient retention and availability, yielding greater crops. Sileshi and colleagues concluded that integrated soil fertility management based on intercropping with N fixation trees decreases vulnerability of crops to drought stress and that the augmented SOC stock in these agroforestry systems promotes soil conservation. A number of additional agroforestry systems that involve fertilizer trees and are widespread throughout Africa were described by Garrity et al. (2010), who revealed an overall improvement in soil fertility, increased crop productivity, and augmented C sequestration in above and below ground biomass. However, as noted by Kandji et al. (2006), utilization of leguminous trees or crops may increase N2O emissions as compared to non N-fertilized agro-ecosystems.

While agroforestry systems contain less C than primary or managed forests, they sequester C over and beyond what would occur under other agricultural activities. In a wide-scale study conducted in Ghana, Wade et al. (2010) compared the C stock of natural forests to that of non-intensive and intensive cocoa (Theobroma cacao) based agroforestry systems. The mean total C sequestration capacity was of 155 Mg C ha−1 for the forests, 131 Mg C ha−1 for the non-intensive systems, and 39 Mg C ha−1 for the intensive systems. In addition, Wade and colleagues observed that decreased cropping intensity resulted in an increase in species richness of the native trees left after forest conversion. Henry et al. (2009) reported strong and positive relationships between tree species diversity and above ground biomass-C in several Kenyan agroforestry systems.

In a recently published synthesis study, Nair et al. (2009) reviewed many works dealing with C sequestration by various agroforestry systems. They summarized that root C stocks can range between 1.3 and 20.5 Mg ha−1 and that shoot C stocks can range between 6.3 and 172.0 Mg ha−1. Nair and colleagues added that the total SOC stocks in these systems may range between 24 and 90 Mg ha−1 in arid and semi-arid lowlands, 21–173 Mg ha−1 in tropical highlands, and 10–235 Mg ha−1 in humid lowlands. In a synthesis study conducted in Costa Rica and Canada, Oelbermann et al. (2004) reported a greater total input of organic matter in tropical than in temperate agroforestry systems. However, the greater input did not necessarily increase the SOC stock when compared to a temperate system of similar age as a result of faster turnover rates. Oelbermann and colleagues estimated that the C sequestration rate in aboveground biomass is 2.1 Pg C year−1 in the tropical biome and 1.9 Pg C year−1 in the temperate biome. Furthermore, they commented that capacity to sequester C varies greatly and depends on a range of physical and biotic conditions, as well as on management practices. Zomer et al. (2009) calculated that agroforestry systems support a total population of ∼1.5 billion people across more than 1 billion ha in developing countries.

The advantageous nature of agroforestry systems has led to their inclusion in the afforestation and reforestation activities that are eligible for payment under the Clean Development Mechanism (CDM, see: UNFCCC 2003). Hence, trading the sequestered C is a viable economic opportunity for practitioners of agroforestry systems, who are primarily subsistence farmers in developing countries (Nair et al. 2010). Measuring total C stocks before the establishment of agroforestry systems and evaluating their dynamics are necessary steps to get approval for funding. Allometric equations that relate the tree’s height and diameter to its biomass allow non-destructive estimates of the above- and below-ground woody stocks (Schoeneberger 2009). Currently, many of the costs associated with developing small holder C offset plantings are related with the need for a large number of field-based measurements. In order to reduce costs, measurements of both biomass-C and SOC should be combined with spatial analyses and relevant information technologies. However, considering the great diversity of agroforestry practices, it is impossible to utilize a single protocol to fit all project types. Further development is needed in advanced methods for quantifying inventory and dynamics of C stocks, as well as in elaborated protocols for C offset in order to allow payments for a wider range of agroforestry projects (Skole et al. 2009). In the USA, several federal and state conservation programs provide financial incentives to landowners for the establishment and management of agroforestry plantings (Schoeneberger 2009). An example is the 2002 Farm Bill (Farm Security and Rural Investment Act of 2002) that has authorized the US Forest Service to allocate 100 million dollars on forestry projects in the private, non-industrial sector. This mechanism functions under the Forest Land Enhancement Program (FLEP) and supports a range of agroforestry practices through several tracks such as cost-sharing, land rental, incentive payments, and technical assistance (USDA 2002).

4 Biochar as soil amendment

Vegetation growth is a very efficient, natural way of CO2 immobilization, since photosynthesis captures more CO2 from the atmosphere than any other natural process. However, the efficiency for long-term C sequestration by biomass is limited because large portion of the C is unstable, returning to the atmosphere as CO2 through decomposition and respiration in a short time of months to years. Therefore, in order to considerably increase long-term C sequestration, biomass has to be converted to a relatively non-degradable form, such as biochar (Lehmann 2007a; Lee et al. 2010).

Biochar is a by-product of the C-negative pyrolysis technology for production of bio-energy from organic materials. The process is conducted under complete or partial exclusion of oxygen and relies on capturing the off-gases from thermal decomposition of organic materials (Lehmann 2007b). Roberts et al. (2010) reported that the net energy of the pyrolysis technology is highly positive, that is, generating much more energy than is consumed. Coupling the pyrolysis process with application of the by-product biochar in soil actually removes CO2 from the atmosphere, as more C is sequestered than that emitted (Roberts et al. 2010). The biochar is highly resistant to microbial activity, considerably augmenting the recalcitrant fraction of SOC and decreasing emissions of CO2 from soil (Glaser et al. 2002; Lehmann 2007a). In addition, biochar application was reported to decrease emissions of CH4 (Rondon et al. 2005), and N2O (Rondon et al. 2005; Yanai et al. 2007; Roberts et al. 2010) from soils. A range of agro-ecosystems in which biochar was applied in their soil are presented in Table 3. Despite representing a wide range of feedstocks and crops, the overall impact is the increase in SOC stock, decrease in emissions of GHGs, lesser leaching of nutrients, and smaller contamination of off-site water sources.

Table 3 Impact of biochar application on emissions of GHGs from soil and additional ecosystem services
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Despite the recalcitrant nature of biochar, about 40% of the total biomass-C of the feedstock is lost during the pyrolysis process, and an additional 10% is mineralized over a few months after biochar application in soil. Nevertheless, the remaining 50% of the total C is relatively stable (Laird et al. 2009). The degree of stability of the biochar-C depends on its specifications. While C in biochar produced by high-temperature pyrolysis is either recalcitrant or degradable at an extremely slow rate, some of the C in biochar produced under low temperatures is biodegradable. In addition, compared with fallow soils, application of biochar increases rates of CO2 emissions from the amended soil. This response may be explained by several factors, such as lower bulk density, improved aeration, and higher pH, providing a favorable habitat for soil microorganisms (Laird et al. 2009).

Biochar application positively impacts soil fertility, for example, through its effect on cation exchange capacity (CEC). Increase in CEC is attributed to the existence of carboxylic groups on the surfaces of the biochar itself, as well as due to exposed carboxylic groups of organic acids absorbed by the biochar, both of which contribute negative surface charge to the biochar (Liang et al. 2006). In addition, biochar can effectively absorb ammonia (NH3), reducing its loss through volatilization. However, the biochar itself contains a limited amount of available mineral nutrients in its ash content, and therefore, its application in soil is usually done in conjunction with fertilizer management (Lee et al. 2010). Nonetheless, the high capacity to absorb nutrients enables their retention in the rooting zone, increasing fertilizer efficiency while decreasing the leaching of nutrients and reducing contamination of underground water sources (Glaser et al. 2002; Laird et al. 2010a). In addition, the high porosity of biochar augments physical quality of the soil, increasing its water-holding capacity and hydraulic conductivity (Laird et al. 2010b). The combined effect of biochar application is the enhancement of soil fertility and productive capacity, resulting in increased crop yields (Marris 2006).

The improved structure formation and stability of the soil are expected to decrease frequency and magnitude of soil loss through erosional processes. However, the impact of biochar on soil erosion has been only very scantly studied and, therefore, requires much more research. In their study, Foereid et al. (2011) modeled the loss of biochar through erosional processes. Foereid and colleagues assumed that larger pieces of biochar are more prone to erosion. They reported positive relations between rate of soil erosion and content of biochar in sediments. In addition, they found that on a 2,000-year time scale, the largest loss (75%) of biochar is through soil erosion, followed by decomposition (20%) and downward movement (0.5%). Overall, it could be expected that retention of biochar on the soil surface with no incorporation into the soil would increase its vulnerability to erosion by either water or wind force. Therefore, also in long-term NT agro-ecosystems, biochar should be incorporated into soil. This could be done by the means of occasional inversion tillage.

According to the International Biochar Initiatives, 154 biochar projects have been carried out in 43 developing countries in 2011. These projects have taken place in a range of operational and functional scales, including household (26%), farm (41%), village (12%), cooperative (11%), and regional (10%) levels. The feedstocks include a range of organic residues and wastes including timber mill waste, plant material derived from invasive species, leaves, livestock manure, plantation prunings, field stover, rice husks, and waste of fruits such as cacao, coffee, and coconut. The biochar production scale varies from small to large and technologies vary from batch retort kiln to continuous-process kiln (Kelpie 2011). Considering an application rate of between 10 and 100 Mg biochar per hectare (Chan et al. 2007) and that biochar’s C concentration is between 50% (Lehmann et al. 2006) and 78% (Gaskin et al. 2008), and assuming a total area of 1,411 M ha cropland around the world (Lee et al. 2010), then the global capacity for storing biochar-C under this landuse is between 7 and 110 Pg. Lee et al. (2010) noted that additional infinite amounts of biochar could be sequestered in abandoned mines, underground reservoirs, and geological formations.

Similar to other bioenergy sectors, feedstocks for the pyrolysis technology can include a range of agricultural and organic wastes, such as forestry residues, prunings from fruit trees, domestic wastes, livestock- and human-derived sludge, and municipal sewage. In addition, designated agricultural crops that do not require high inputs in water and fertilizer and that could be grown on marginal lands that cannot be used for food production could also be utilized for bioenergy crops (Tilman et al. 2009). At the same time, residues of agricultural crops, such as wheat straw or corn stalk, must not be considered as viable biofuel feedstocks. This is because the retention of residues on the soil maintains the SOC stock, supports the soil food chain, protects the soil surface from raindrop impact and erosional processes, preserves the soil quality, and sustains the productive capacity of the agro-ecosystem (Lal and Pimentel 2009).

The basic concept of biochar production is simple, and therefore, it can be easily produced locally by farmers, including those with low economic power (Glaser et al. 2002). Some examples for low-cost means for production of charcoal were detailed by the FAO (1987), describing the construction of earth pits, earth mounds, brick kilns, and metal kilns. Another low-cost technique is the two-barrel kiln that can be constructed by using reused materials (Fig. 4), which can be used to produce biochar from a range of agricultural wastes (Fig. 5). Such a low-cost kiln has been used during the last year in the Dead Sea & Arava Science Center of Israel in order to demonstrate the ease of biochar production. The rate of producing biochar in this kiln has been approximately 5 kg from ground date palm fronds during a 1-h process and about 20 kg from cattle manure during a 4-h process. Actually, the simplicity of the kiln and its low-cost enable the running of up to five kilns simultaneously by one person, reaching a maximal biochar production capacity of about 25 kg from fronds over 1 h, or 100 kg from manure over 4 h.

Fig. 4
figure 4

A home-made kiln, produced of reused barrels. An ∼80-l barrel, filled in with ground date palm fronds is laid upside down within an ∼160-l barrel. The gap between the barrels is filled with (unground) fronds that are used as fuel. About 5 kg biochar are produced during a 1-h process. Using the same method, ∼20 kg biochar are produced from cattle manure during a 4-h process. Notice the piles of dry date palm fronds in the background, acquired from the date palm orchard (in the far background). Picture was taken in the Arava Valley, southern Israel

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Fig. 5
figure 5

Biochar made of ground date palm fronds. Produced using the home-made kiln shown in Fig. 4. Picture was taken at the Dead Sea & Arava Science Center, Israel

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However, it may be noteworthy to mention that in many parts of the world, several socio-cultural obstacles prevent a much wider expansion of the biochar management practice. Of the main obstacles, gender barriers (Kelpie 2011) and land tenure issues are especially remarkable. In addition, there are several gaps in understanding various aspects related to the production process, as well as to the application in soil. For example, much research is needed to quantify the interactions among biochar, soil properties, management practices, and emissions of CO2 and N2O (Laird et al. 2009). In addition, the potential impacts on soil quality and crop yields under a range of soil’s nutritional stati, as well as under alkaline soil conditions, are yet unclear. Furthermore, the production of biochar from many types of feedstocks and under a wide range of pyrolysis processes makes validations rather difficult. Therefore, standardization of biochar is crucial to enable comparisons among the different feedstocks and production procedures.

5 Regulations

Until the late 1990s, the CO2 produced during burning of fossil or woody fuels was considered valueless, and the cheapest means to dispose it was merely to let it be emitted into the atmosphere. In the same manner, the increased oxidation of SOC caused by the conversion of native ecosystems into cultivated lands was not treated and considered as a legitimate environmental cost of agricultural production. As long as the emission of GHGs is free of any costs, the economic rationale of the relevant sectors is to increase agricultural production with the associated increased emissions of CO2. Likewise, emissions of CH4 through cultivation of rice paddies and animal husbandry, as well as emissions of N2O through fertilizers use, are expected to increase.

Considering that the capacity of the atmosphere to absorb GHGs prior to the emergence of catastrophic climatic changes is limited, some actions are inevitable (IPCC 2007; Lal 2007). For example, relying more on renewable (solar/wind/geothermal) and other alternative (biomass-derived) energy sources, enhancing efficiency of energy use, and increasing C sequestration (Kyoto Protocol 1998). Large rates of C sequestration may be achieved by means of vegetative biomass (van Minnen et al. 2008) and soils (Lal et al. 2007). Nevertheless, as stressed by Hardin (1968) in his essay, “The Tragedy of the Commons”, which deals with the (mis-) management of land resources, leaning exclusively on technical solutions without relying on change in human behavior and values, might not suffice. Therefore, education and sharing of knowledge, along with the use of incentive-based means, are crucial. Indeed, history has shown that economic incentives are very effective in reducing emissions of GHGs. Since the 1960s, the only reliefs from global gas emissions have corresponded with economic downturns or high energy prices. Similarly, C taxation is expected to reduce demands, encouraging individuals, companies, and countries to decrease energy consumption or to replace high C-emission fuels with more efficient ones (Meyerson 1998). Currently, coupling the impacts of rising energy prices and climate change has already had a dramatic impact. Food prices are increasing sharply along with the price of consumer products and services, threatening the economic wellbeing of billions of people. The current steep rise in oil prices boosts inflation, which has long been regarded as the most unequal “tax” as it redistributes wealth from low income to high income groups. It is virtually certain that these circumstances will worsen in the years ahead, endangering the Western world’s vision of a wealthier and more equitable society (Carvalho et al. 2011).

The inter-governmental consideration of C as a valuable commodity under the Kyoto Protocol (1998) has initiated the required shift in the willingness and motivation of individuals, private sectors, and governments towards responsible behavior. The CDM described under Article 12 in this protocol allows developed countries, which have been committed to GHGs reduction to purchase emission credits by investing in mitigation projects in developing countries (Nair et al. 2010). Under this mechanism, developing countries may generate Certified Emissions Reductions, which can be sold to committed developed countries. This mechanism aimed at stimulating clean development in developing countries by providing finance for environmentally friendly technologies, while developed countries gain access to low-cost mitigation projects (Whitman and Lehmann 2009). Afforestation and reforestation are among the eligible means under the CDM, aimed at increasing above- and below-ground biomass production, as well as at accumulating plant residues and sequestering SOC (UNFCCC 2003). Agroforestry has been also included under this framework, providing farmers in developing countries with attractive economic benefits (Nair et al. 2010). However, considering the current market price for C, a wide implementation of agroforestry projects seems unfeasible due to the small size of many farms that do not achieve a critical land area to enable compensation for the concomitant costs. Higher financial compensations for C sequestration in agroforestry projects, which at the same time improve ecosystem services, would allow clearer win–win scenarios for small holder farmers, encouraging them to implement these practices (Henry et al. 2009).

An example for the mode of operation of the financial bodies may be the case of the BioCarbon Fund, launched by the World Bank, and aimed at funding “projects that sequester or conserve C in forests and agro-ecosystems” in order to “deliver cost-effective emission reductions, while promoting biodiversity conservation and poverty alleviation.” Despite that “the BioCarbon Fund can consider purchasing C from a variety of land use[s] and forestry projects,” the only eligible practices are “afforestation and reforestation, [and] reducing emissions from deforestation and [land] degradation.” Nonetheless, the BioCarbon Fund “is exploring innovative approaches to [sequester] agricultural C” (BioCarbon Fund’s website: http://wbcarbonfinance.org/Router.cfm?Page=BioCF).

To date, except for agroforestry, other agricultural landuses and management practices are not eligible for funding through the CDM. Even the promising practice of biochar application in soil is not eligible for payment under this framework (Marris 2006; Whitman and Lehmann 2009). Thereby, an inter-governmental intervention should adjust the global C finance market’s framework to include the biochar management practice. The incentive benefits should be high enough to promote the implementation of this practice by farmers. For example, Roberts et al. (2010) calculated that valuing an Mg of CO2 at a minimum of ∼$60 would encourage farmers in the USA to pyrolyze switchgrass for co-production of bio-energy and biochar. In addition, a considerable increase is necessary in authorizational payments for maintaining ecosystem services such as soil erosion control, water quality preservation, and biodiversity conservation. This would promote the adoption of such environmentally friendly farming practices, aimed at sustaining global food security and alleviating poverty while reducing environmental footprint and minimizing GHGs emissions.

6 Conclusions

The impact of agriculture in emissions of GHGs is enormous. Intensively managed croplands have emitted considerable amounts of CO2, CH4, and N2O, tremendously increasing their atmospheric concentrations. Furthermore, some of the intensive farming practices have increased soil erosion as well as contamination of off-site water sources. At the same time, conservation agriculture practices may decrease emissions of GHGs, mitigate climate change, on the environment. However, as emphasized in this review, the agronomic efficiency of some of the conservation practices are site-dependent and not relevant to all geographic zones and climatic conditions. Nevertheless, specific conservation practices—agroforestry systems and application of biochar in soil—can boost sequestration of organic C and, in addition, increase fertilizer efficiency, enhance productive capacity, and advance global food security. At the same time, these practices support a range of ecosystem services such as decreased soil erosion, reduced contamination of off-site water sources and increased species diversity and ecosystem health, and therefore, can be utilized in reclamation of degraded lands. Future regulations should facilitate national and international schemes of payments for these agricultural practices, encouraging their wide implementation throughout the world.