by Kathleen Draper

Make carbon sink, not rise. Most of the world’s toilets are not connected to sewers but drain into septic systems or cesspools. With new smart biochar based septic designs, this human derived waste carbon could become long term carbon sinks. Refashioning our subsoil waste disposal into safe systems of carbon sequestration could open up vast new opportunities for the circular urban economy and climate action.

The need to find large reservoirs for carbon storage is heating up exponentially. While some Carbon Dioxide Removal (CDR) technologies are targeting deep crevices into which CO₂ can be injected, biochar has historically focused on additions just below our feet, particularly in carbon-depleted or contaminated agricultural soils. Few that deeply understand biochar would argue against its use on farms, though more than a few argue against its use anywhere else! However, getting farmers to adopt new practices is notoriously challenging and slow, even when the value proposition looks favorable.

The stark climate reality is that we do not have the luxury of time on our side to solely focus on persuading vast numbers of farmers to bury carbon. The biochar industry has spent more than a decade focusing on farmers and has barely put a dent in that market. This is not to say we should give up on it. It is obvious that we cannot and should not. But if biochar is to have a material impact on climate change in the short time frame that is necessary, vast new sinks must be identified quickly.

Image 1: Scheme of a conventional sanitary sewer pipe trench.

Fortunately, there are a growing number of not only safe but beneficial sinks that could provide enormous sequestration opportunities. [Albert Bates and I explore many of these in BURN.] One such opportunity could be upgrading the lowly cesspool, a primitive system used to manage toilet waste, as well as the improved but far from optimal septic system. Cesspools, though largely now banned in the US and Europe, continue to exist in rural areas around the world and cause significant issues as they are often a primary source of Nitrogen pollution and usually contain coliform bacteria (e.g. E.coli), phosphates, chlorides, chemicals and viruses (e.g. the novel Corona virus may spread through human feces) which contaminate groundwater or local water bodies.

Cesspools or cesspits come in various types but are basically a drywell for untreated household waste, including human feces. In some rural areas and even suburban areas these are still common. Long Island still has 250,000 cesspools and 110,000 septic systems in need of upgrading or replacement as these are largely responsible for the increase in algal blooms, fish kills and beach closures. Local municipalities have offered significant funds to homeowners to upgrade these leaching liabilities which have harmed the local economy.

The renewal or replacement of cesspools and septic systems may become a huge opportunity for the establishment of carbon-sinks. Septic systems are perhaps an even greater carbon sink opportunity as they are far more common and typically much larger. Though they are an improvement over cesspits, they too are responsible for significant water contamination.

A properly designed septic system often involves moving around large quantities of soil, sand and gravel to improve infiltration and filtration. However, septic systems seem to work indirectly on the principal of ‘dilution is the solution’ versus actual filtration. Soils with poor percolation are amended with vast amounts of sand and gravel to improve infiltration of rain water. The majority of this filtering materials is, strange as it sounds, placed above the area where the effluent is released into the soil. This set-up allows rainwater to percolate and prevents effluent from backing up and pooling at the surface. Drain pipes with holes at the bottom are typically sandwiched into a gravel layer which sits directly atop the subsoil (to where the effluent is then diluted and may contaminate ground water and even surface water). Adding a filtration layer of biochar below the drain pipes could help immobilize or neutralize contaminants of concern.

Using biochar for a septic system for private construction

In perhaps the first residential use of biochar in a septic system, I recently worked with a civil engineer and excavator to design a shallow trench septic system that included biochar. Since biochar is not currently an approved soil amendment for drain fields, we had to design it so that it could still pass regulatory muster. This was done by digging deeper trenches upon which biochar could be added before the gravel, sand and topsoil. Before trenches were dug, sand was used to create a raised bed. In this design roughly 10 cm of biochar was applied to 5 trenches measuring 18m in length for a total of one ton of biochar. Drainpipes with holes near the bottom were sandwiched in between 30 cm of gravel then covered with a geotextile to prevent soil from clogging the gravel. Finally, 30cm of topsoil was layered above the geotextile. Adding a layer of biochar beneath the gravel and pipe enhances adsorption and biological degradation, though in this demonstration measuring results will be quite a challenge.

Image 2&3: Setup of the septic biochar system where the biochar bed is placed below the sewage drainage pipe. Sand and gravel is then placed above.

This was a small system yet it still required 8 truckloads of sand (200t) and 2 truckloads of gravel to be hauled in. Larger homes often required three times as much sand in areas with heavy clay soils.  Mixing far larger quantities of biochar into the sand and/or gravel layer, might allow for 50 – 100 tons of biochar per septic system. In addition to the drain fields, it could be used around the septic tank and in the trenches connecting the septic tank to the home and to the distribution box which lies between the septic tank and the drain field lines and serves to distribute effluent evenly. Similar structured soils have been successfully used in the Stockholm Biochar Project. Storm drains, bioswales, retention basics, detention ponds, rain gardens and dry creeks could also benefit from the addition of biochar. Biochar could even be used as an additive to make the concrete septic tank, the PVC pipes and the geotextile used in the septic system!

Municipalities could potentially fund cesspool upgrades or out of date septic systems using carbon financing. Imagine replicating the Stockholm model where the city converts green waste into biochar. The municipality could sell the carbon removal credits and use both the funds and the biochar to upgrade the leaking systems. This would have the added advantages of creating large-scale, local demand for biochar while also helping cities to manage their organic materials more economically.

Reusing septic system amendments

Municipalities might consider retaining the rights to the amendments so that they could remove them after a decade or so and use the enriched structured soils to rebuild coast lines or for urban tree planting. In theory, the biochar septic system could be set-up like a sort of “night soil aging” though the economics would be hard to predict and municipalities would be content, probably, with the assessment of the added carbon sink.

Carbon Math

The biochar for my homebuilding demonstration project was supplied by National Carbon Technologies using sustainably sourced wood and had a carbon content >90%. One ton of biochar would have 900kg of carbon. Based on experience from calculations using the EBC-Sink methodology, an average carbon to CO₂ multiplier of 2.5 could be used to convert the carbon to CO2e.  This takes into consideration the emissions related to feedstock production, chipping, transportation, pyrolysis emissions and other factors. The current carbon removal price is north of $60 per ton CO₂e, so this single ton could generate $US 150 per ton of biochar. Now imagine being able to put 50 – 100 tons per residential system. That would generate more than enough funding to install higher performing, healthier sewage management systems. In Long Island alone this could lead to 18 – 36 Million tons of biochar being utilized which equates to carbon sequestration of between 45 Mt – 90 Mt CO₂eq! Now imagine this model being rolled out across a good part of the planet. This is the scale that is needed to make a material impact on climate rebalancing.

Additional research needed

To date little, if any, formal research has been done on this topic. Understanding the role of biochar’s particle size, friability, water holding capacity and adsorption potential would be helpful to understand if certain biochars are superior to others in this context and how much is needed for an average household of six persons. Determining what blends of biochar, sand and gravel work best in different soils and climates would help in developing best management practices needed for regulators and system designers. Designing a cost-effective circular economy model for carbonizing urban green waste and blending it into reusable structured soils for cesspools and septic systems would be a game-changer when it comes to creating multi-beneficial sinks.

The post Septic Systems and Cesspools as C-sinks appeared first on Regen Agri-Source.

by Kathleen Draper

Organic residuals from wastewater treatment may just be one of the world’s most renewable, yet underutilized organic waste streams.  One of the most common end uses, land-application, is increasingly restricted so communities across the globe are seeking alternative management practices.  Carbonizing this waste stream may solve many of the problems associated with biosolids management while also offering the possibility of new ways to decarbonize an industry which contributes nearly 3% of global GHG emissions.

The challenge of managing wastewater residuals

Treating waste water from residential, commercial and industrial sources as well as storm water poses enormous challenges and costs to the municipalities that operate waste water treatment plants (WWTP). Sewage sludge, a by-product of waste water treatment, contains both beneficial organic macro nutrients such as nitrogen and phosphorus as well as micronutrients.   It also contains harmful elements derived from solvents, pesticides, beauty care products, cleaning products, and pharmaceuticals, which get flushed down toilets and drains.  Some of these can be minimized through different stabilization processes, but other elements persist. The public is increasingly vocal about how and where residuals from WWTPs are applied due to odors and emerging constituents of concern in sludge and biosolids.  Regulators are also increasing standards for biosolids recycling in an effort to reduce contamination of water sources and soils. This has sent many WWTP operators on a quest to find improved stabilization and disposal options. Carbonization of biosolids may be a key technology which could provide many advantages over current practices. 

Current Practices for Wastewater Residual Management

WWTP operators have a handful of options for the management of sewage including land application, landfilling, composting, lime stabilization, incineration, and anaerobic digestion.  Benefits and drawbacks of the most common methods are discussed below.

Land Application

The oldest practice known to man for disposal of human waste is land application.  Traditionally land application has been one of the most common (55% in the US) and lowest cost disposal methods for sludge or biosolids although this is largely dependent on proximity to where sludge is spread. Utilizing the macro and micro nutrients in sludge has been a cost-effective way for farmers to reduce the need for fertilizers.  However public perception and regulatory oversite has instigated a shift away from this disposal method at least as far as application to land used for growing food.  Other factors including application rates, seasonal restrictions, pH of the waste and nutrient management aimed at protecting watersheds restrict how, when and where biosolids can be applied.

Although some developing world countries have few regulations when it comes to land application of biosolids, many countries in Europe forbid it. Regulations in Europe and elsewhere stipulate strict heavy metal thresholds and nutrient loading rates when it comes to land application of any soil amendments .  However certain materials found in biosolids are currently unregulated, and in many cases untracked, that are emerging as constituents of concern.  These include PCBs, pathogens (e.g. fecal coliforms, salmonella), bacteria (e.g. E.coli), solvents, dioxins, pesticides, microbeads, microfibers, and heavy metals (e.g. copper, nickel, zinc, etc.).  Some of these contaminants persist throughout the waste water treatment cycle and sludge stabilization process. Sludge which is treated is often referred to as ‘biosolids’ to differentiate it from untreated sludge.  Once applied to soils, these contaminants can find their way into plants and ground water as well as both land and aquatic animals leading to long term negative impacts on entire eco-systems if not properly managed.

Landfill

Many WWTP operators send their residuals to landfills as they are unable to find other outlets for it.  In the U.S. more than 2 million tons of sludge/biosolids are sent to landfills (NEBRA 2004). This disposal method is both costly and comes with a heavy carbon footprint due not only to the transportation of residuals but also due to fugitive methane emissions related to decomposing biosolids. The extremely high moisture content of sludge and biosolids creates some challenges for landfills in terms of spreading solids, and odors can be a concern.  The quality of landfill leachate is also negatively impacted by the pathogens and heavy metals found in biosolids. 

Incineration

In Europe roughly 21% of WWTP dispose of sludge via incineration (Kellessidis et al., 2012) while less than 200 of nearly 16,000 facilities in the US do so.  Incineration can reduce residuals by 70 – 95% (US DOT, NYDEC). The resulting sludge ash is most often landfilled but has found application in some areas as a filler for cement, for use in road construction or as daily landfill cover.   While energy is required during the start-up phase of incineration, the heat produced becomes largely self-sustaining for most of the processing period. Concern over emissions can often be mitigated by emission control technologies.  Although incineration of sludge has a lower carbon footprint than land filling, valuable nutrients (N, P) are wasted. It should be noted that there are emerging technologies capable of recovery P from ash which are becoming mandatory in some places.

Anaerobic Digestion

Anaerobic digestion of sewage sludge provides many environmental and financial benefits.  The overall amount of residuals is reduced by up to 55% (Wong et al., 2011) which translates into reduced transportation costs and tipping fees for landfilling biosolids as well as reduced wear and tear on roads. This can be a significant savings to a plant’s bottom line as the cost of landfilling continues to rise.  Odors are reduced, though not eliminated and the thermal conditioning which occurs within the AD reduces though does not completely eradicate certain elements of concern.  Perhaps most importantly, the biogas generated reduces the amount of electricity that municipalities must purchase from the grid to run the WWTPs.  This can be a considerable cost savings as the energy usage from WWTPs can often account for 40% or more of the total energy budget for some cities. 

Despite all of these benefits, AD technology is still not prevalent in many parts of the world due to steep up-front capital costs and the current low cost of natural gas.  In the US the percentage of WWTP’s with ADs is estimated at less than 10% (Lono-Batura et al., 2012).  While Europe has more than 10,000 ADs converting organic waste into biogas, few are used for digesting sewage sludge. 

Carbonization can improve management of residuals

Carbonization of biosolids could provide substantial benefits to WWTPs. Although relatively few examples currently exist for pyrolyzing or gasifying biosolids, R&D activities are increasing. In Europe the EU funded “PyroChar” a multi-year, multi-stakeholder project aimed at designing all needed components for pyrolyzing biosolids for small communities.

By-products of thermal conversion of sewage sludge include synthetic gas (syngas), bio-oil and pyrolysate (the chared solid). The benefits of carbonized biosolids are discussed below. The benefits and uses of the other byproducts is beyond the scope of this article.

Carbonized sewage/biosolids

While carbonized biosolids contain valuable nutrients such as phosphorus, they contain relatively low levels of carbon as compared to other feedstocks used for biochar production.  Given the low carbon content, carbonized biosolids may or may not be considered as biochar depending on the standards used for classifying biochar.  Under the International Biochar Initiatives (IBI) Biochar Standards, if all other properties fall within acceptable ranges, biosolids char could be considered as a Class III biochar (IBI 2011).  The European Biochar Certificate (EBC, 2012) Standards have a higher threshold for carbon content. To be considered as biochar, chars must have at least 50% carbon.  For chars below this amount, they are denoted as ‘pyrolysis ash containing biochar’ or ‘pyrolysates’.

The actual amounts of carbon and other elements in sewage derived char will vary considerably depending on the production temperature as can be seen in the table below (Hussein et al., 2011). 

  TABLE 1: Proximate, ultimate and agronomic properties of wastewater sludge & sludge biochar

Benefits of Carbonized biosolids

GHG reductions

The amount of greenhouse gas (GHG) emissions which result from WWTP are estimated to account for 2.8% of total global GHG emissions. WWTPs use enormous amounts of energy to process wastewater causing large CO2 emissions.  Methane caused by degrading organic material and N2O caused by degradation of nitrogen components in the wastewater (e.g., urea) can be considerable. Although these can be significant, they are quite variable depending on the disposal method used.  Landfilling generates the highest emissions and land-application the lowest (Miller-Robbie et al., 2015).

Converting biosolids to char eliminates CH4 emissions related to landfilled residuals as no further decomposition would occur.  Further reductions related to reduced transportation of residuals would also improve the carbon footprint of WWTPs.  Given that the carbon in many biochars is fairly recalcitrant, landfilling or land application could represent a carbon sequestration opportunity, though further research is needed to understand and measure longevity of this particular type of char in different soils and in landfills. Carbonized biosolids would also greatly reduce the amount of leachate produced at landfills which can be costly to manage. Biochar will also adsorb CH4 emitted from other organic material in landfills.  Thus carbonized biosolids could be particularly useful to use when closing landfills that do not have methane capture technology in place. 

Toxin reductions

Triclosan (TCS) and Triclocarbon (TCC) are antimicrobials commonly found in toothpaste, soap, shampoo, mouthwash and pesticides.  These elements persist in sludge and biosolids and are toxic to certain aquatic organisms. Nonylphenol (NP) is a chemical compound frequently found in laundry and liquid detergents which is an endocrine disruptor.  TCS, TCC and NP are not currently regulated but are commonly found in sewage sludge and are contaminants of emerging concern if sludge is applied to land.  Pyrolysis can eliminate certain antimicrobials and sulfactants (Ross et al., 2015) when the correct temperatures and residence times are used.  The length of time sludge spends in the reactor also influences contaminant reduction as can be seen in Figure 1.  

Fig. 1 Impact of pyrolysis reaction time on removal of TCS, TCC, and NP from biosolids during pyrolysis at 500C. Source: Ross et al., 2015

Another persistent micro-pollutant is estrogen which is found in medicines (e.g. birth control) and livestock waste.  Estrogen has been found in biosolids and anaerobically digested biosolids and has been found to have negative impacts such as causing intersexing of aquatic animals.  Hoffman et al. (2006) found that pyrolysis at temperatures above 400C could reduce estrogen levels in sludge by >95%. With higher pyrolysis temperatures estrogen levels and other organic pharmaceuticals can be eliminated though more research is needed to define the pryrolysis treatment conditions necessary to obtain these results for selected substances. 

Reduced disposal costs

Hussain et al. (2010) found that pyrolysis can reduce sewage volume by nearly half at higher temperatures.  Tipping fees and transportation of biosolids can be substantial for WWTPs.  For example, one small WWTP in upstate NY spends nearly $300,000 per year to landfill 4,500 tons of biosolids per year. So even if no markets were to be found for the biosolids char, the cost savings from reduced disposal costs could be substantial.

Odor mitigation

Nuisance odors occur throughout most of the wastewater treatment process including during thickening, digestion, dewatering, storage, truck loading, transportation, drying, composting, etc. (US EPA 2000).  As micro-organisms break down the amino acids and carbohydrates found in sewage sulfur, ammonia and other malodorous compounds are emitted.  Such odors impact quality of life and property values and have led to an increase in local ordinances that ban land application of biosolids. While the odors during the initial processing stages would not be impacted, carbonizing sludge and biosolids would eliminate this problem completely at disposal sites.

Markets for Carbonized biosolids

Soil application

Pyrolysates made from biosolids could be land applied much as biosolids currently are but without the negative impacts.  The thermal conversion process effectively reduces certain constituents of concern yet retains at least part of the nutrients and carbon.  Phosphorus in biosolids char is to a good extent plant available and could render the char into a type of slow release fertilizer while also reducing the risk of leaching (Bridle et al., 2004). While a good portion of N is retained, Bridle et al. (2004) found it to be insoluble at least in the short term.

Table 2 Nutrient levels in sludge and char, source: Bridle et al., 2004

The carbon in the char is more recalcitrant than carbon in biosolids and could potentially help improve soil carbon levels over the long term. As can be seen in table 1, the pH of biosolids is acidic whereas the pH of biochar ranges widely depending on production temperatures.  It is feasible that WWTPs could customize the pH in biosolids char for particular land application needs.  Using different production parameters and post-processing technologies it is possible to create a range of pH from a low pH near 5 up to a char with a high pH of more than 10. While certain heavy metals would be retained in the char in a more concentrated form, they would be less mobile than they are in sewage sludge (Liu et al., 2013). 

Using the biochar classification scheme developed by Camps-Arbestain et al. (2015) a 50/50 mix of pyrolyzed biosolids and wood chips would be a class 2 (mid-range) fertilizer due to P and Mg content; a class 4 (highest class) liming agent and a class 1 (lowest class) carbon storage material.

Applying a dryer material such as charred biosolids would likely be easier on farm equipment than applying higher moisture content biosolids.  It would also be safer for farm workers in terms of eliminating the risk of pathogen contamination. Odors would also be eliminated which would be a welcome change for all neighboring communities where biosolids are currently applied. 

Some studies (Hossein et al., 2010) have found that sludge char can improve yields, especially when combined with fertilizers.  Hossein et al. (2014) found that fertilizer enhanced sludge char showed an increase of 167% in cherry tomato production compared to a non-fertilized control, 27% over the fertilized control and 60% over the non-fertilized sludge char. They assessed bioavailability of heavy metals and most were well below permitted maximums with the exception of cadmium which was at the maximum allowable amount in Australia. They also showed, as can be seen in Table 3, that increased production temperatures generally, though not always, tended to increase concentrations of certain micro-nutrients and heavy metals.

Table 3: Means for certain nutrient concentrations & metals of wastewater sludge and biochars. Source: Hussein et al., 2010

However Van Wesenbeeck et al. (2014) showed that certain pyrolyzed sludge chars retain metals that are above permitted levels in some areas and therefore cannot be used for land application.  They found that some metals are released during thermal conditioning (e.g. mercury) while others were retained.  In their studies zinc, molybdenum and chromium were above allowable levels in Hawaii and cadmium, copper, nickel and zinc were above maximums permitted in Belgium.  

Co-product for power plants

Using low temperature carbonization technology, a WWTP in Hiroshima, Japan creates a high calorie, low odor char which is sold to a coal fire plant resulting in an estimated reduction of 6,400 tons of CO2 (Matsumiya 2012).  Another plant in Tokyo has been producing 8,700 tons of sewage sludge char from 99,000 tons of sewage which is then mixed with coal for energy generation.  The sewage is heated at 500C for 1 hour and generates 2,000 kcal kg–1 of heat calorie, which is equivalent to one third of that of coal (Yachigo et al., 2013).

A US company, PHG Energy, installed a gasifier two years ago that carbonizes biosolids mixed with wood waste (70% wood waste, 30% sludge).  The resulting 12t per day of charred materials is used to co-fire electricity plants as well. They expect to go live with a second project in the fall of 2016 which will process 64 t per day and produce 34 t per day of char.

Filtration medium

Char from biosolids could potentially be used as a filtration medium particularly those produced at high temperatures.  Increased temperature has been found to increase surface area and adsorption properties. Rio et al., (2015) found that sludge char and limed sludge chars could effectively reduce acid and basic dyes and phenol via filtration.  Researchers in China found that a blended char of sludge and tea waste could effectively adsorb certain pollutants such as methylene blue, a dye (Fan et al., 2016). Chen et al., (2002) found that pretreating anaerobically digested sewage sludge with a reagent and then pyrolyzing it at 500C created a char with “remarkable micropore and mesopore surface areas and notable adsorption capacities” for certain contaminants.

Reclamation Projects

In the quest for large scale, inexpensive disposal methods, biosolids have been used for reclamation projects to help revegetate mine lands and mitigate erosion and leaching.  Fly ash or lime is often added to improve pH and make application of biosolids easier.  In Pennsylvania the use of biosolids for reclamation is broadly used.  Although other organic substrates might be far more beneficial, biosolids enabled revegetation which helped reestablish wildlife.  However some studies have shown that high biosolids application rates negatively impacted water quality.  Increases in acidity levels that resulted from biosolids application are linked to increased leachability of Al, Mn, Cu, Ni, Pb and Zn (PennState, 2016).

The low C/N ratio, low pH and odor issues related to the use of biosolids could be moderated if combined with char.  The use of charred sludge in lieu of biosolids in reclamation projects would reduce the risks of heavy metal and inorganic N leaching, and it can be expected that soil remediation would be improved using the adsorbing matrix of sewage pyrolysates.

Building materials

Lack of suitable land for application of sludge and fewer landfills that accept sludge has spurred recent research on alternative uses for biosolids including its use in construction materials.  Research shows that combining sludge with concrete or sintering it with clay is a viable method of stabilizing heavy metals. (REF)  Japan has created a growing market for the use of dewatered sludge as a raw material in the production of Portland cement.  Researchers in Brazil have found that small amounts of sewage sludge could effectively be incorporated in roof tiles (Ingunza et al., 2015).  Taiwanese researchers baked sewage sludge at 250C and found that using up to 10% torrified sludge could still produce good quality bricks (Weng et al., 2013).

Biochar has also been found to serve as a viable building material and provides various benefits from improved humidity control to insulation (see here in tBJ).  Charring sludge would not only reduce volumes, but could provide improved properties for construction materials.  It could also provide a possible cascading use scenario, where sludge is charred at WWTPs, then used as a filtration medium at the facility and finally encapsulated in building materials.

Conclusion

As the world’s population continues to grow, so too will the amount of biosolids that need to be managed. Carbonizing this unending supply of biomass could reduce many, if not most, of the negative environmental impacts associated with biosolids: odors, toxic leaching, GHG emissions, etc.  In addition charred biosolids could generate a range of new products which could displace current products which have high carbon footprints such as activated carbons used in filtration or light-weight concrete which can safely sequester carbon.

Technology to carbonize high moisture feedstocks such as biosolids has been available for several years but has yet to catch on in most areas of the world.  As the cost of organics disposal and the cost of carbon pollution increases, the economics of carbonizing biosolids will inevitably become much more attractive.

References

• Bridle, T.R. and Pritchard, D., (2004). Energy and nutrient recovery from sewage sludge via pyrolysis; ResearchGate – Water Science & Technology
• EBC, 2012. European Biochar Certificate – Guidelines for a Sustainable Production of Biochar. Version 7.1 of 22th December 2015 [WWW Document]. Eur. Biochar Found. URL http://www.european-biochar.org/en/download (accessed 1.12.16).
• Fan, Shisuo, Tang, Jie, Wang, Yi, Li, Hui, Zhang, Hao, Tang, Jun, Wang, Shen, Li, Xuede, (2016) Biochar prepared from co-pyrolysis of municipal sewage sludge and tea waste for the adsorption of methylene blue from aqueous solutions: Kinetics, isotherm, thermodynamic and mechanism, Journal of Molecular Liquids, Volume 220, August 2016, Pages 432–441
• Hoffman, T.C., Zitomer, D.H., McNamara, P.J., (2016) Pyrolysis of Wastewater Biosolids Significantly Reduces Estrogenicity; Journal of Hazardous Materials, Volume 317, Pages 579–584; http://dx.doi.org/10.1016/j.jhazmat.2016.05.088
• Hossein, Mustafa K., Strezova, Vladimir, Chanb, K. Yin, Nelson, Peter F., (2010) Agronomic properties of wastewater sludge biochar and bioavailability of metals in production of cherry tomato (Lycopersicon esculentum), Elsevier – Chemosphere 78 (2010) 1167–1171
• Ingunza, Durante,  Del Pilar, Maria, Andressa Dantas, Lima, (2015) Use of Sewage Sludge as Raw Material in the Manufacture of Roofs, International Conference on Civil, Materials and Environmental Sciences
• Rio, S., Faur-Brasquet, C., Le Coq, L., Le Cloirec, P., 2005. Structure Characterization and Adsorption Properties of Pyrolyzed Sewage Sludge. Environ. Sci. Technol. 39, 4249–4257.
• Liu, Taoze, Liu, Bangyu, Zhang, Wei, Nutrients and Heavy Metals in Biochar Produced by Sewage Sludge Pyrolysis: Its Application in Soil Amendment, Pol. J. Environ. Stud. Vol. 23, No. 1 (2014), 271-275
• Lono-Batura, Maile, Qi, Yinan and Beecher, Ned; Biogas Production And Potential From U.S. Wastewater Treatment, BioCycle December 2012, Vol. 53, No. 12, p. 46
• Matsumiya, Yosuke, (2012). Green Energy Production from Municipal Sewage Sludge in Japan Japan Sewage Works Association http://gcus.jp/wp/wp-content/uploads/2014/06/ebd9e233be72625b03c96047573177f9.pdf
• Miller-Robbie^,, Leslie, Bridget A. Ulrich^a, Dotti F. Ramey^a, Kathryn S. Spencer^b, Skuyler P. Herzog^a, Tzahi Y. Cath^a, Jennifer R. Stokes^c, Christopher P. Higgins, Life cycle energy and greenhouse gas assessment of the co-production of biosolids and biochar for land application; Journal of Cleaner Production; Volume 91, 15 March 2015, Pages 118–127, 
• PennState Department of Ecosystem Science and Management, Use of Biosolids for Mine Reclamation: Assessment of Impacts on Acid Mine Drainage and Nutrient Discharge, 2016; accessed 19^th August 2016:  http://ecosystems.psu.edu/research/labs/land-analysis/projects/state/use-of-biosolids-for-mine-reclamation-assessment-of-impacts-on-acid-mine-drainage-and-nutrient-discharge
• Ross, John Fate of Micropollutants During Pyrolysis of Biosolids 2014 http://epublications.marquette.edu/cgi/viewcontent.cgi?article=1283&context=theses_open
• U.S. EPA (2000) Biosolids and Residual Management Fact Sheet: Odor Control in Biosolids Managemen
• Van Wesenbeeck, Sam, Prins, Wolter ,  Ronsse, Frederik,   AntalJr., Michael Jerry, Sewage Sludge Carbonization for Biochar Applications. Fate of Heavy Metals Energy Fuels, 2014, 28 (8), pp 5318–5326
• Weng , Chih-Huang , Lin , Deng-Fong , Chian, Pen-Chi (2003) Utilization of sludge as brick materials; Elsevier – Advances in Environmental Research pp 679–685
• Yachigo, Mieko, Sato, Shinjiro  (2013) Leachability and Vegetable Absorption of Heavy Metals from Sewage Sludge Biochar INTECH – Soil Processes and Current Trends in Quality Assessment Ch 15, pp  399 – 416

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by Henning Gerlach & Hans-Peter Schmidt

The poultry industry is struggling more and more with livestock disease. Often this can be traced back to microbial pathogens and ammonia in the litter. The addition of highly porous biochar can serve to reduce toxic ammonia pollution in the coops and regulate the moisture level of the litter. The biting coop odour and foot pad dermatitis in the poultry can be prevented within just a few days. If biochar is included in the feed, toxins can be deactivated already in the digestive system. The intestinal flora is positively activated, and the vitality of the animals improves rapidly and markedly.

Industrial poultry farming places extremely high demands on hygiene of the coops, of the air in the coops and of the feed, as well as of waste and faecal matter. High animal densities increase the pathogen pressure as the immune response of stressed animals is weakened, with the result that more pathogens are excreted. The smaller the area in which the animals are kept, the more the microbial environment in the coop is dominated by microbes that live off the animal itself and its excretions. This produces a significant risk of spreading germs, which is exacerbated by poor coop and feeding hygiene.

If, in addition, the poultry is treated with anti-infection and antibacterial agents, this creates an environment that selects pathogens that are resistant to the drugs being used. Because these events depend on the quantity of pathogens (pathogen pressure), it is all the more important to control the coop environment in a timely manner so that pathogen pressure is reduced.

Due to the loose housing of poultry, animals in coop systems inevitably live in constant contact with their excrement. The extremely nutrient-rich and humid faeces create ideal conditions for the multiplication of pathogenic microorganisms. Added to this, the microbial decomposition of the excrement leads to significant emissions of ammonia. The pungent-smelling gas is harmful to the animals because it irritates the mucous membranes, attacks the lungs, weakens the immune system and even accumulates in the blood. Besides the effect on animal welfare, animal performance also deteriorates seriously. Last but not least, ammonia emissions are environmentally harmful. Via nitrogen return in rain, they produce highly climate-damaging emissions of nitrous oxide, soil acidification and eutrophication of water bodies.

The use of biochar as a feed additive and as litter amendment can significantly minimize the problems described both with regard to animal health and in terms of environmental performance.

Instructions for the use of biochar in litter

Biochar has a very high water holding capacity and can absorb up to 5 times its own weight of water. Biochar adsorbs very efficiently both organic molecules such as amino acids, fatty acids, proteins and urea and also mineral compounds such as ammonium, ammonia and nitrate. Used in litter, biochar locks in moisture and organic and inorganic nitrogen compounds. The nitrogen adsorption and the continuous drying of the litter deprive the microbial pathogens of their nutrient base and reduce toxic emissions of ammonia. After just a few days, a significant reduction in coop odour can already be noticed.

With the lowering of the moisture content and ammonia contamination the risk of footpad diseases decreases. Existing infections begin to heal. Animals’ resistance increases, with a positive effect on their vitality, egg production and final body weight.

Biochar’s high adsorption capacity makes it possible to reduce the use of lime in the litter, thereby reducing the pH of the litter and manure, which in turn reduces ammonia emissions.

Footpad diseases

Turkeys and broilers frequently suffer from leg weakness syndrome, which last but not least is economically disastrous. To this should be added footpad inflammation, known as footpad dermatitis (pododermatitis). The causes of these inflammation reactions are multifactorial, but the main causes are high levels of ammonia (NH3) and overly damp litter. Particularly important in this respect are the structure and hardness of the litter, both of which are improved by the use of biochar.

The effects of footpad diseases include pain, reduced physical activity, reduced feed and water intake, growth depression, feather pecking/cannibalism, reduced carcass quality and increased mortality.

Application of biochar

The biochar should, depending on the type of litter, be mixed 5-10 vol % with the usual litter. The char is first moistened in order to prevent dust formation. Ideally it is applied in the form of lactic acid biochar bokashi. When using straw pellets as litter, the char is best added already at the pelleting stage.

If silage is used as litter, the char can already be added at the ensiling stage. In this way, dust formation can be avoided entirely, and the low pH of the silage kills off pathogens. Mixed into silage, the char is bound very well and no longer rubs off onto the animals’ feet. This is particularly important in egg farms, since coal can rub off from the hens’ feet onto the eggs.

Use of biochar in feed

In addition to its use as a litter additive, biochar, and in particular biochar bokashi, is also used as a feed supplement. Biochar promotes digestion, improves feed efficiency, and thus in particular energy absorption via the feed. Toxins such as dioxin, glyphosate, mycotoxins, pesticides and PAHs are efficiently bound by the biochar, thereby obviating any adverse effects on the digestive system and intestinal flora. The health, activity and balance of the animals will also be improved, as will meat and egg production. With animals’ immune systems stabilized, the risk of infection from pathogenic micro-organisms decreases.

The huge economic impact of diarrhoeal diseases in poultry is well-known. The causes of these diseases are often of an infectious nature and are caused by, among others, E. coli, clostridia, coccidia and mycobacteria. Of particular importance are salmonella and campylobacter germs; while rarely causing disease in poultry, they can do so in humans. Non-infectious causes of disease are in particular poor feed quality and biocide contamination of the feed, as when herbicides are used to siccate feed grain or to treat weeds during the growing of GMO corn or soy feed. The consequences are an increased susceptibility to disease, growth depression, infertility and digestive disorders.

Numerous factors are responsible for the stabilization of the intestinal milieu. Of particular importance here are the stabilization of the intestinal barrier and the functionality of the liver. Numerous bacteria such as lactobacilli and enterococci, but also non-pathogenic yeasts play an indispensable role here. Feeding biochar and biochar bokashi can stimulate the activity of these desired microorganisms in the digestive system. The benefit of the biochar lies therefore not least in its ability to relieve in particular the liver-intestinal circuit.

The charging of the biochar with specific lactobacilli to direct the symbiosis in the gastro-intestinal tract of farm animals can further potentiate the effect of the biochar. Biochar bokashis produced as ready-made feed on the basis of a fermented biochar, wheat bran and herbs are an important feed supplement for maintaining and enhancing performance in animal production.

According to studies by Van (2006), the addition of up to 0.6% biochar in the feed improves growth in young animals by an average of 17%. Similar results are confirmed by Kana (2010) and Ruttanvut (2009) for ducks and broilers. No systematic scientific studies of long-term effects exist as yet.

It is recommended to mix 0.4% – 0.6% biochar to the usual feed. With laying hens the feed supplement should be suspended for 2-3 days every 10-15 days. Biochar bokashis, such as Carbon-Feed from Swiss Biochar, should be added 2% – 3%% to the usual feed. If biochar is already used in the feed, the amount of biochar in the litter can be reduced accordingly.

Using biochar to improve manure quality

The above-mentioned effects of biochar for storing moisture and nutrients also mean that the poultry manure is better degraded microbiologically. Carbon and nitrogen losses are significantly reduced and with them the emission of greenhouse gases (Steiner 2010). The fertilizer quality of the poultry manure increases strongly as a result of the biochar and the odour pollution can be reduced significantly, which increases the marketing potential of poultry manure.

If biochar is used neither in the litter nor in the feed, it is advisable to sprinkle it in a ratio of 10 vol % on the manure belt.

If the poultry manure is used for energy production in biogas units, the addition of biochar both increases the methane yield and improves the fertilizing quality of the digestate. Poultry manure can also be directly pyrolyzed to produce biochar and energy.

Literature

Kana, JR, Teguia, A, Mungfu, BM, Tchoumboue, J 2010, ‘Growth performance and carcass characteristics of broiler chickens fed diets supplemented with graded levels of charcoal from maize cob or seed of Canarium schweinfurthii Engl’, Tropical Animal Health and Production 43(1):51–56.
Steiner C, Das KC, Melear N, Lakly D, Reducing nitrogen loss during poultry litter composting using biochar, J Environ Qual. 2010 Jul-Aug; 39(4): 1236-42
Van, DTT, Mui, NT & Ledin, I. 2006, ‘Effect of method of processing foliage of Acacia mangium and inclusion of bamboo charcoal in the diet on performance of growing goats’, Animal Feed Science and Technology 130: 242-56.

Ithaca comments

Certainly, the use of biochar can partially mitigate the catastrophic situation in mass animal husbandry. Infectious diseases are slightly reduced, feed intake slightly improved, meat weight increases and there are less greenhouse gases. But do we really want to help making this outrageous contempt for the life of mass-farmed animals even more efficient? The statement that the animals would thus suffer less should be seen as bitter sarcasm.
Should we not expect that the properties of biochar will be used precisely to allow the use of poorer and even more contaminated feed, because the char fixes the toxins? Feed manufacturers will certainly very soon be mixing the biochar directly into their pellets, making it almost impossible to measure dioxins and PAHs any longer with the traditional analysis tools. Will not the better coop environment following the use of biochar in the litter be used precisely in order to increase animal density even further and build the sheds even larger?
In the corn pellets with which the chickens are fed, the level of glyphosate (RoundUp herbicide) is allowed to be 200 times higher than in other foods. This herbicide is eliminated not only via the droppings, but especially via the eggs, of which the highly bred hens lay one every day. However, the usual analytical methods for the detection of herbicides in foods work only in low-protein foods. It is not even possible to estimate the magnitude of herbicide contamination in the eggs…
Is it right for us as ecologists and veterinarians to propagate solutions that extend the continuance of this farming system, even if the environmental damage is reduced in the medium term? Can we talk our way out of this by saying that factory farming is a societal problem that can only be solved by society? I really do not know. I only can attempt to publish the information as completely as possible and from different points of view.
But one thing I do know: healthy chickens that run free in green fields and have access to uncontaminated feed would need biochar only in exceptional cases as a remedy. All the diseases that we write about in the article would be eliminated with a free-range grazing system, as shown in the following movie, and this even without biochar. Should the first step in the above-mentioned dilemma not be rather to eat eggs from chickens that are allowed to run daily in the grass?”https://www.youtube.com/embed/nx9R5Hv0KN8

On the manure belt of this chickenmobile biochar makes sense for enhancing the fertilizing properties of manure in the overall system. If we think further such innovative models and ideas, other unimagined possibilities such as a combination of vegetable, fruit and wine growing with poultry farming open themselves up.

The post Biochar in poultry farming appeared first on Regen Agri-Source.

by Hans-Peter Schmidt

One of our oldest preconceptions is that a cowshed inevitably stinks. But the pungent odour of liquid manure is first and foremost the sign of a microbial decomposition process that has gone out of control. That which stinks to high heaven is not only an offence to delicate citizens’ noses but above all a source of disease for the animals living there. Thanks to biochar and to the control of the microbial environment in the shed and in the manure pit, materials cycles can be closed. Liquid manure in this way becomes a highly efficient, sustainable and odourless fertilizer.

In spring and autumn, when farmers spray their fields with liquid manure, an acrid stench spreads across the countryside. This pungent smell comes especially from ammonia, a volatile nitrogen compound formed from the urea contained in the manure. Large portions of the ammonia, which is corrosive to soil organisms and fine roots, escape into the atmosphere, where it binds to dust particles and returns in the form of acid rain onto fields, forests, cities and water systems, causing major environmental damage.

While some of the minerals in the liquid manure like ammonium, nitrate, urea and phosphate become available as nutrients to the plant, a, significant portion of the nutrients is leached to ground and surface waters, not to mention the climate-damaging gas emissions. Some 50% of the nitrogen is lost by degassing, leaching and erosion between cowshed and field via the manure pit, resulting not only in very low fertilizer efficiency but costly environmental damage. In Germany alone, agricultural ammonia emissions amount to more than 600 000 tons a year.

Due to the outgassing of ammonia and leaching of nitrates, fertilizing with untreated liquid manure results in soil acidification, which in turn greatly impairs the fertility and biological activity of the soil and accelerates the decomposition of humus.

Even more dangerous than soil acidification, however, is when, as a result of using non-treated manure, pathogens from animals’ digestive tracts or from the bacteria strains and fungal spores generated in the rotting manure pit are spread into the fields. Although most disease-causing microorganisms are destroyed by antagonists in the soil, some highly resistant strains of bacteria, fungal spores and other pathogens like Clostridium (EHEC, botulism) survive the entire plant growth cycle and can be ingested by the animals via fodder from these fields. This completes a vicious cycle that gradually breeds more and more resistant pathogens and may endanger both animal and human health.

Modern liquid manure spreaders can considerably decrease nutrient losses and greenhouse gas emissions. But the highly concentrated ammonia still remains harmful to the soil. Pre-treating liquid manure and mixing with biochar may increase fertilizer efficiency by around 50%.

What stinks in the liquid manure?

Most of the nitrogen comes into the manure in the form of urea, which is then converted into ammonia and CO2 (or to ammonium + CO3) through the enzyme urease. Urease is found in animals’ stomachs, in the manure pit and in the soil; which means that the decomposition of urea can take place virtually continuously and everywhere. While conversion in the soil into the important plant nutrient, ammonium, is a desired process and the key to the fertilizing effect, conversion in the manure pit means the loss of the nitrogen and, not least, the acrid stench.

Manure treatment by lactic acid fermentation

Disinfecting the liquid manure, stabilizing the nutrients and thereby preventing the development of odour require:

1. binding the volatile nutrients,
2. inhibiting the enzyme urease, and
3. preventing the proliferation of pathogenic microorganisms.

Lactic acid fermentation, a very old method of preservation, is particularly suited for stabilizing the urea in the manure and preventing it from rotting. During lactic acid fermentation, sugar compounds, which occur in all plant materials and also, in lower concentrations, in liquid manure, are converted by lactic acid bacteria into lactic acid. This conversion lowers the pH down to 3.5 – 4.5, resulting in a highly acidic environment below the survival conditions for most types of bacteria, spore-forming organisms and enzymes.

The prevention of rotting and outgassing is not solely a question of the acid, otherwise one could use any acid to lower the pH. It is of great importance that the acidic environment is produced by lactic acid bacteria, since in this way the remaining sugar compounds in the liquid manure are broken down, depriving competing putrefactive bacteria of the nutrient base and preventing them from multiplying further.

In addition, useful cell components such as nitrogen, phosphorus, sulphur and carbon are stored in the cells of the rapidly proliferating lactic acid bacteria. Nitrogen, phosphorus and sulphur that are stored in the cell tissues of lactic acid bacteria are biologically tied down and no longer volatile. Competing microbes find themselves thus in short supply of all essential nutrients, in an unfavorable, acidic environment.

When the lactic acid bacteria subsequently come with the stabilized liquid manure onto agricultural land, they will in turn be deprived of their survival conditions by the atmospheric oxygen and the higher pH values of the soil, whereby the nutrients stored in their cells will be recycled by other microbes and become available to plants. The result is a truly biologically activated fertilizer.

Historical significance of lactic acid fermentation

Lactic acid fermentation was used already in the Stone Age for preserving food. This was a precondition for the development of a stockpiling economy, which in turn permitted the formation of sedentary societies. Lactic acid fermentation was used to produce sauerkraut, sourdough bread, yogurt, sausage, wine and also for feed silage. It was, however, hardly used for fermenting liquid manure treatment, as in traditional cowsheds no liquid manure was produced and solid manures are better composted than fermented.

Instructions for treating liquid manure with sauerkraut juice and biochar

In order to make use of lactic acid fermentation to ‘conserve’ liquid manure, the manure pit must first be inoculated with a sufficient quantity of lactic acid bacteria. The lactic acid bacteria have to be multiplied in such a way that pH falls below 5. The following procedure has proven successful in practice:

1. Empty the liquid manure pit, leaving not more than 25 cm of sediment.
2. Inoculate the remaining manure in the pit with 0.2 – 0.5% sauerkraut juice.
3. To encourage the multiplication of the lactic acid bacteria, add 1% molasses.
4. Mix in 2% biochar for fixing nutrients and toxic substances.

(for 50 m3 of manure sediment this amounts to 100-250 l sauerkraut juice, 500 l molasses and 1 m3 biochar)

Sauerkraut juice contains a very high quantity of lactic acid bacteria and is ideal for inoculation. Instead of sauerkraut juice one can also use bread drink (Brottrunk), silage juice, or EM-A (effective microorganisms). The latter also contains other microorganisms which favorably influence the process. Sauerkraut juice, however, is by far the cheapest means and of assured quality. Today, millions of litres of sauerkraut juice are disposed of, against payment, in wastewater treatment plants. The agronomic use of sauerkraut juice would also be a good example for the closing of nutrient cycles.

Molasses are needed so that the lactic acid bacteria in the sauerkraut juice can multiply thousands of times and so optimally adjust the microbial environment of the liquid manure. If, however, the manure pit is too full, it will not be possible to shift the microbial environment from putrefying to lactic acid fermentation, because the other microbial strains will be too dominant and the lactic acid bacteria will not prevail in spite of the molasses. It is therefore absolutely necessary to pump the pit as empty as possible before initiating the lactic acid fermentation.

Once the pH of the liquid manure drops below 5 through the effect of lactic acid fermentation, the conversion of urea to ammonia is prevented, rot-causing bacteria are suppressed, the liquid manure no longer stinks and the nutrients are retained and fixed.

Effect of biochar

The high specific surface area and the high cation exchange capacity of biochar makes it very efficient in binding ammonium and ammonia and other odourous and often toxic substances. For this reason the use of biochar, even without lactic acid fermentation, is rapidly effective. Through the use of biochar most of the nitrogen in the manure can be stored available for plants. The leaching of manure nutrients in the soil is slowed significantly, which not only protects the groundwater, but also and in particular prevents the acidification of the soil. Biochar-treated liquid manure promotes soil activity and humus formation. Soils are built up over the longer term instead of being eluviated by toxic liquid manures. Overall, the fertilizer efficiency of liquid manure may nearly doubling with biochar.

Although the treatment of manure with biochar is already effective on its own, its combination with lactic acid fermentation as described above is still recommended, as it is only through lactic acid fermentation that the liquid manure is disinfected and pathogenic bacteria are killed off.

Manure treatment during the year

To maintain the microbial environment in the manure pit, the liquid manure added each day has to be inoculated with lactic acid bacteria. This can be done by adding approximately 0.1% sauerkraut juice and biochar by volume to the daily addition of fresh manure into the pit. The most effective way of doing is is spraying the sauerkraut juice directly in the cowshed using an automatic atomizing system. This ensures a healthy microbial environment already in the cowshed and is good not only for the manure but especially also for the livestock and for the people working in the shed.

Treating slurry with biochar on the Holderstock farm. (Photo: Wilhelmine and Bruno Koller)

The climate in the cowshed will change already after a few days. It will stop smelling unpleasant. Also, the cattle will be noticeably calmer, and inflamed udders and hooves will cease swelling. Atomizing in the cowshed can be done completely automatically, with 2 litres of sauerkraut juice atomized every four hours per 100 livestock units.

Biochar in silage and feeds

Biochar too should be used in the cowshed as early as possible. Biochar can already be introduced into the feed silaging process. Biochar promotes lactic acid fermentation of the silage and prevents faulty fermentation. Less acetic acid and, in particular, less butyric acid are formed during the silaging process, reducing the risk of clostridia infections. The risk of fungus formation in the silage and of the related mycotoxins is also reduced. Biochar has a very high water holding capacity, which ensures good fermentation quality and prevents the production of butyric acid, especially in cases where the silage does not have enough time to wilt properly ( (e.g. bad weather) and at excessively low dry weight concentrations.

Overall, preservation quality is improved by moisture buffering. Thanks to biochar, there is little or no formation of fermentation juices, that are to be feared owing to the formation of butyric acid.

Biochar fixes pesticides and heavy metals that come with the biomass into the silage, where both negatively affect the fermentation environment and subsequently have a toxic effect on the animals. When present in the finished silage, biochar improves digestion and increases the energy conversion from the feed.

In using biochar in silage, only biochars registered as food additives and produced by licensed feed manufacturers may be used. Plants and charcoal have been used as feed additives ever since the early Iron Age to regulate digestion, in particular in ruminants. Biochar increases feed efficiency, which in turn increases the animals’ energy performance. Toxic substances are bound and the microbial environment in the digestive tract is stabilized. Biochar certified as feed can be added directly into the feed at 0.5%. To prevent the risk of blocking essential nutrients, feeding of biochar should be interrupted every 14 days for at least 5 days.

Summary

The combined use of biochar and lactic acid bacteria leads to improved animal health and production capacity, stabilizes the climate in the cowshed, disinfects it and the liquid manure, prevents nutrient losses and greenhouse gas emissions and leads finally to a biologically efficient fertilization of farmland. Biochar is not a panacea, but by supplementing good farming practice can sustainably optimize processes, increase farming profitability and make a positive contribution to the environment.

The post Treating liquid manure with biochar appeared first on Regen Agri-Source.

by Hans-Peter Schmidt, Samuel Abiven, Nikolas Hagemann and Johannes Meyer zu Drewer

An executive summary for Global Biochar Carbon Sink certification

Biochar that was produced at pyrolysis temperatures above 550°C and presenting a molar H:C ratio below 0.4 is highly persistent when applied to the soil. 75% of such biochar carbon consists of stable polycyclic aromatic carbon and will persist after soil application for more than 1000 years, independent of the soil type and climate. The remaining 25% of the biochar carbon may be considered semi-persistent, presenting a mean residence time in soil of 50 to 100 years, depending on soil type and climate. Soil-applied biochar contains thus two distinct carbon pools with different degrees of permanence and therefore has two different carbon sink values. The climate service obtained from the stable fraction of biochar (75% of the C-content) can be considered of equal permanence as geological storage (e.g., DACCS, BECCS, Enhanced Weathering).

Biochar is a heterogenous carbonaceous material (EBC, 2012) that consists of two distinct carbon pools with different degrees of persistence when applied to soil:

• The persistent aromatic carbon (PAC) pool, which consists of larger clusters of aromatic carbon rings (see figure 1), generally with more than seven aromatic rings, is not susceptible to degradation. The PAC pool has a mean residence time (MRT) in soil largely exceeding 1000 years (Bowring et al., 2022; Howell et al., 2022), independent of common environmental factors such as soil humidity, temperature, freeze-thaw-cycles, and biological activity or agricultural practices like tillage. Cluster sizes of aromatic carbons (how the carbon ring structures cling together) may be more important for persistence than the sheer number of aromatic rings in a molecule (Mao et al., 2012; McBeath & Smernik, 2009; Nguyen et al., 2010).

• The semi-persistent carbon (SPC) pool, which contains aliphatic, small aromatic, and heteroaromatic carbon species, is more degradable in soil (Rombolà et al., 2016). Some compounds of the semi-persistent carbon pool can degrade within the first year after soil application; others will persist for decades and even centuries depending on the chemistry of the aliphatic and small aromatic compounds and their physical placement within the porous structure of the biochar. On average, the SPC fraction of biochar has an MRT in the order of at least 50 to 100 years, depending on the biochar composition (i.e., distribution of aliphatic carbon, small aromatics, heteroaromatics), the soil type, and the climate (Bowring et al., 2022; Hilscher & Knicker, 2011; Lehmann et al., 2015; Pisani et al., 2014; Schmidt et al., 2011; Singh et al., 2012). With an MRT of 49 years for soil organic carbon (Schmidt et al., 2011), considering the evidence that pyrogenic carbon persists longer in soil than soil organic carbon (S. Lutfalla et al., 2017; Schmidt et al., 2011) and the calculated MRT of 91 years for 49% of soil-applied pyrogenic carbon in six field trials with none optimized biochars, the application of an MRT of 50 years for the semi-persistent carbon pool of biochar with an H:C < 0.4 is conservative and presents a considerable margin.

In the environment, each of the carbonaceous compounds separated in those two respective carbon pools shows distinct degradation dynamics that can be described by an individual degradation curve. If biochar is incubated for one or two, or even eight years, as done in controlled incubation laboratory studies (Kuzyakov et al., 2014; Lehmann et al., 2015) and the resulting degradation data are then mathematically extrapolated into the far future, the prediction of the degradation dynamic is erroneous because it assumes that (1) the biochar consists only of semi-persistent carbon pools (aliphatics and small clusters of aromatic and heteroaromatic rings) and (2) the exponential biological degradation model is valid, despite degradation mechanisms being rather physico-chemical than biological. Over a time scale of several thousands of years, persistent aromatic carbon moieties may eventually also be degraded (Bowring et al., 2022), but this fact is barely contained in degradation data measured only during the first decade after biochar was incubated in soil.

Figure 1: Schematic representation of different molecular forms of carbon in biochar.

All biochar incubation studies observed that the rate of degradation slows down exponentially with time and that the experimental data can be fitted mathematically with bi- or trimodal decay functions (Lehmann et al., 2015; Wang et al., 2016; Zimmerman & Gao, 2013). However, such mathematical fitting may be misleading as it cannot account for qualitative transitions (e.g., removal of physical protection of compounds inside the biochar structure) occurring decades and centuries after the latest measured data points. As shown by Lutfalla et al. (2019), the small number of existing data sets presenting decadal degradation data of carbon in soil cannot be fitted by such bi- or trimodal decay functions. Therefore, projecting the degradation behavior of the semi-persistent carbon pool onto the degradation curve of the entire biochar is not adequate and biases our understanding of long-term carbon dynamics.

The percentage of PAC in a given biochar depends mainly on the pyrolysis conditions (i.e., temperature, residence time, heating rate, particle size, carrier gas, pressure) but also on the feedstock characteristics (i.e., lignin and ash content of biomass) (McDonald-Wharry, 2021). The PAC content can be quantified by hydrogen pyrolysis (HyPy) (Ascough et al., 2009; Rombolà et al., 2016) or by Raman spectroscopy (McDonald-Wharry, 2021; McDonald-Wharry et al., 2013). HyPy analysis is reliable and proven but too complex to be utilized in commercial labs and thus too expensive to be used in routine analysis as for example, in the EBC certification process. Raman spectroscopy as well as the newer Mid-Infrared or Rock-Eval methods, are more cost-efficient analytical methods and are currently under methodological evaluation for the EBC.

Other parameters of biochar, such as production conditions (e.g., temperature) and elemental composition (i.e., using molar H:C and O:C ratio), which are frequently used as proxies for the degree of aromatization rely on poorly constrained databases with low analytical quality. The dataset used e.g., by Woolf et al. (2021) and the IPCC (IPCC, 2019) to calculate biochar persistence based on the highest treatment temperature (HTT) and H:C proxies, contained only one biochar series where HTT was actually measured (Budai et al., 2014) all other temperatures were estimates. Pyrolysis temperature, i.e., the actual temperature inside the biomass particle during conversion, cannot be measured accurately in most industrial pyrolyzers and does not capture the effects that heating rate, residence time, particle size, and pressure have on the formation of PAC (Santín et al., 2017). Moreover, when using literature data on H:C to parameterize mathematical functions on biochar persistence, more care needs to be taken to only use correctly analyzed carbon and hydrogen contents of materials that are really biochar. The IPCC (2019) and Woolf (2021) included e.g. many materials with H:C ratios above 0.7 which are clearly not biochars (EBC, 2012; IBI, 2015) and considered implausible data resulting from insufficiently described analyses (e.g., how was the biochar dried before measuring the H-content). While the molar H:C ratio can be measured with sufficient precision, some fractions of biochars with low H:C are not necessarily PAC and may be considered semi-persistent only (Howell et al., 2022). The H:C ratio is thus a proxy with limited significance for the quantification of biochar persistence.

Since PAC is not yet analyzed for every EBC-certified biochar, average biochar data from the literature should be used with caution. Conservative security margins must be used to estimate the persistent biochar content and, thus, the portion of biochar carbon that will endure as a C-sink for more than 1000 years. Based on the degradation experiments published so far and considering that the calculated decay functions express only the degradation dynamic of the semi-persistent biochar carbon pool, the carbon calculated as remaining after 100 years is regarded as the minimum PAC fraction of biochars with an H:C ratio < 0.4. Applying the conventionally assumed average degradation rate of 0.3% per year, the 74% of carbon remaining after 100 years (Camps-Arbestain et al., 2015; H.-P. Schmidt et al., 2020) can be considered PAC. This corresponds well to the experimental data presented by Howell et al. (2022), finding 75% stable polycyclic aromatic carbon for various engineered biochars with H:C ratios below 0.4 using the HyPy quantification method. Biochars with H:C ratios above 0.4 are likely to have a distinct labile carbon pool of incomplete pyrolyzed biomass or condensates that may become subject to more rapid biological degradation. Dedicated research for those materials is needed to assess the persistence of pyrolyzed biomass presenting high H:C ratios above 0.4 (Pulcher et al., 2022).

Fig. 2: Sequestration curve of a 1000 tons carbon sink made from soil-applied biochar with an H:C ratio below 0.4. The persistent aromatic carbon (PAC) pool presents 750 t carbon that will remain over more than 1000 years in the terrestrial system. The semi-persistent carbon (SPC) pool has a minimum MRT of 50 years and was modeled on a bi-modal exponential decay function. The complete SPC decay occurs over 350 years. Thus, the total carbon sink decreases to 87.5% after 50 years and reaches the stable PAC plateau of 75.0% of total pyrogenic carbon after 350 years. The decay function is
〖Total PyC〗_((x))=a*e^((-kf*x) )+b*e^((-ks*x) )+P
with a = 45.423, kf = 0.513, b = 212.007, ks = 0.009448, P = 742.5 nd x = year after soil application. The decay curve of the semi-persistent carbon pool is an approximation covering multiple discrete (physical) degradation events rather than a continually harmonious decomposition.

To the best of the current scientific knowledge, it is safe to assume that biochar with an H:C ratio below 0.4 can be best described by a 2-pool-model presenting

1. A persistent aromatic carbon (PAC) pool of 75% with an MRT of >1000 (1400-14400 years, see annex), competitive with geological carbon storage and suitable for CO2-emission compensation and
2. A semi-persistent carbon (SPC) pool with an MRT of at least 50 years offering an additional, valuable climate cooling service, yet of a different quality than persistent aromatic carbon (PAC).

Annex 1

Mean residence times of natural pyrogenic carbon

Calculations of global inputs and deposits of naturally produced pyrogenic carbons (PyC) can attest to how robust and conservative the assumption of these average persistence rates over 100 years is. Forest, bush, and steppe fires are examples of incomplete combustion, which transform part of the biomass into chars, i.e., PyC. According to recent surveys of natural fires, 5-15% of the biomass carbon involved in the fire is converted to PyC (Santín et al., 2016). Natural PyCs are similar in structure and material properties to industrially produced biochar. However, it can be assumed that the stability of high HTT industrial biochar, and thus the mean residence time (MRT), is even higher than that of natural PyC (Howell et al., 2022; Santín et al., 2017) due to more controlled and homogeneous high-temperature conditions.

Mainly through forest and steppe fires, about 0.114-0.383 Gt (Giga tons) of pyrogenic carbon (PyC) are generated each year (Santín et al., 2016). Globally, the total mass of PyC in soils is 71-212 Gt, in nearshore sediments 400-1200 Gt, and in further ocean sediments 80-240 Gt (Bird et al., 2015; Santín et al., 2016), resulting in a global PyC pool of 550-1,650 Gt (excluding PyC in water bodies and groundwater sediments). Based on the dimension of the global PyC pool and the annual input of PyC of 0.114 – 0.383 Gt given above, the average MRT of natural PyC can be calculated as

MRT = (GlobalPyC-pool)/annualPyCinput

The MRT range of natural PyC could thus be calculated as (550 Gt / 0.383 Gt a-1 to 1,650 Gt / 0.114 Gt a-1 =) 1,440 to 14,500 years. This time frame is confirmed by Bowring et al. (2022), who determined a minimum MRT of 2,760 years using the same data basis but without including sedimentary PyC.

If we use the extrapolation of Reisser et al. (2016), according to which the PyC content of soil organic carbon (SOC) is 14%, and the global content of SOC is 1,500 to 3,000 Gt (Scharlemann et al., 2014) the global PyC content in soils would be about 210 – 420 Gt (Leifeld et al., 2018). From the annual PyC input of 0.114 – 0.383 Gt, the MRT for PyC in soils would be (210 Gt / 0.382 Gt a-1 to 420 Gt / 0.114 Gt a-1) 550 to 3,700 years. Since the MRT of PyC in sediments is significantly higher than in soils, the difference between the two calculations is plausible. Note, however, that most of the PyC in nearshore sediments is originally derived from PyC leached from soils (Coppola & Druffel, 2016), so that much longer MRTs than the calculated 550 to 3,700 years would result for soil-PyC, except that the pyrogenic carbon would no longer be found in soils but as deposits in sediments (Coppola et al., 2014).

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by Bjorn Embrén

Urban trees face various challenges which frequently lead to high tree mortality, shorter lifespans and increased maintenance cost. To improve tree health and survivability the Swedish capital Stockholm has been testing and refining the use of structured soils and biochar for nearly 10 years. These structured soils consist of gravel mixed with smaller soil amendments such as peat, sand, clay, lava and more with great success: biochar.  In some cases 6 year old trees planted in structured soils with biochar were five times larger than 30 year old trees planted using more traditional urban tree planting techniques. 

Over the past years, a variety of different soil amendments, biochar & stone mixtures have been tested for planting urban and suburban trees, bushes, perennials and grasses in Stockholm.  The Biochar Journal discussed the various experiences and recommendations with Bjorn Embrén who heads up the City of Stockholm’s Landscaping Department. 

Urban trees

Pavements add significant constraints to a tree’s ability to thrive as they severely limit the availability of water and oxygen to tree roots.  Suffocation and dying trees are a common result of these constraints.  Bjorn Embrén and his team have developed a combination of strategies, including the use of structured soils, to effectively re-create a more natural environment for urban trees. The goal has been to recreate, as closely as possible, nature’s hydrological pathways and filtration mechanisms. 

The first step was to replace the heavily compacted soils underneath the pavement with a thick layer of stones as stones are impervious to compaction which enabled the exchange of gases and penetration of water.  Initially two separate layers of different sized rocks were used, but experience has shown that using a deep layer of stones that range in size from 32 – 63mm works equally well if not better.

Fig. 1: Preparing the biochar bed for urban tree allee plantation in Stockholm. 

In the early years of using structured soils, the team flushed soil down between the stones to provide sufficient growing media for the trees.  Based on their earlier successes with suburban tree experiments, the Stockholm team switched, since 2009, to using biochar in lieu of soil for all of their urban tree planting projects.  They have found that the main advantage of biochar for urban tree projects over other soil amendments is that it resists compression and compaction, which is seen as one of the biggest threats for trees and other perennials in urban parks and streets.  The crushed stones and biochar mix results in an improved void ratio, also known as porosity, in the soil (roughly 40%).  Increased porosity facilitates better gas exchange, permeability leading to improved root penetration. 

Fig. 2: Stockholm scheme of urban tree planting.

A further improvement on their methods has been to have the biochar and stones delivered pre-mixed, looking a lot like coal.  They have found that this saves up to 25% in time and labour costs.  The preferred biochar particle size of the biochar is between 1 – 10 mm and currently they recommend approximately 2.25 cubic meters of biochar (2,250 litres or 594 gallons) for each urban tree.  Embrén and his team have been experimenting with different amounts of biochar and have tested a range of 10% – 25% by volume to understand the best balance between water holding capacity (WHC) and infiltration flow rates. They have also been experimenting with different types of fertilizers to embed in the char to promote long term health in urban trees. Having biochar pre-fertilized for optimal urban tree growth would provide significant advantages to urban landscaping teams.

A concrete bunker or planting box is placed atop the large stones to house and protect the tree roots. This has the added benefit of preventing tree roots from penetrating or clogging underground pipes as well as damaging the pavement surrounding the tree. Smaller sized crushed rock is added after the tree is planted.  Critical to preventing this carefully designed aerated environment from getting clogged is the addition of geotextile fabric just beneath the pavement.

Suburban trees

While suburban trees don’t have to deal with pavement stress, they still suffer from compaction and other issues.  Based on his decades of experience growing orchids using biochar (charcoal), Embrén decided to experiment with the use of biochar to try to improve tree survival.  Biochar is much more porous than sand or clay which helps to improve water holding capacity but it also doesn’t biodegrade or compact like some of the other amendments they had been using such as peat. Scouting about for a source of biochar, he eventually found charcoal which was considered by the manufacturer to be ‘bad’ but only because the moisture content was high.  (This would in fact, be a positive benefit for Embrén’s intended use!) 

Understanding that trees cannot grow on stones alone, a thin layer of soil mixed with biochar was added atop a layer of stones.  Traditionally a soil layer of 30 cm is used, but for their early trials they used a mere 10 cm of the 50/50 (by volume) soil/biochar mix. After these initial trials they have used less biochar and have achieved similar highly satisfactory results.  They have also switched to using smaller stones which can more easily be turned with a spade.

Another highly successful technique which has been used to plant 20,000 cherry trees is depicted in Figure 3.  This 70 cm deep and 15 cm wide trench or French drain can be quickly dug with the right equipment allowing for quick yet effective planting. Pre-mixed char and gravel can then be backfilled and saplings planted.  It is likely that the amount of biochar used for these types of plantings will increase, but that may partially depend on the availability and price of biochar. 

Fig. 3: 20 000 cherry trees planted in biochar and chrushed stone. Scheme of a French drain to plant suburban allees.  

On average trees planted in structured soils with biochar have grown approximately 1 meter every year – something which was unheard of in Stockholm.  The grass which grows around these trees is thick and lush.  Embrén has also surprised to see mycorrhiza developing in the biochar enriched substrates, something which he hadn’t previously seen anywhere else in Stockholm.

Parks, perennials and planting beds

While urban and suburban trees have gotten much of the attention, trials in parks, planting beds and round-abouts have also been happening.  Recipes vary depending on the situation but a common and successful mix for perennials and bushes is a blend of 3 parts gravel (2 – 6 mm in size) to 1 part biochar.  Trees seem to require less biochar so a blend of 85% gravel (32 – 63 mm in size) to 15% biochar is used.

Fig. 4: Stockholm roundabout. Substrate: (1) bottom layer with crushed stones (32 – 63 mm) with 15% biochar and fertilizer. (2) top 30 cm layer with small crushed stones (2 – 6 mm) and 25% biochar and fertilizer.  

On one large roundabout of 2000 square meters (see figure 4), the substrate mix contained 15% char mixed with stones (32 – 63 cm). In the top 300 mm layer smaller stones (2 – 6 mm) were used and 25% fertilized biochar.  An unexpected benefit appears to be that traffic noise seems to be lower around the round-abouts where the structured soils with biochar have been used.  One hypothesis is that the increase in subterranean voids may be absorbing sound waves.

Overall according to Embrén results have been so positive that they would like to incorporate biochar into all urban, suburban and perennial projects.  He estimates that doing so would require roughly 800 tons of biochar annually.  Not only would this improve the life span of the many urban plantings and reduce costs associated with replacing trees, but it could create a vast carbon sink with the ability to sequester about 2,000 tons of CO_2^e.

Storm Water Filtration

Based on experience using activated carbon in aquariums as a filtration medium, the team was interested to see if biochar would help filter rain water. They therefore began to add a base layer consisting of pure biochar to act as a filter for contaminants and have been quite pleased with their observations. Although they don’t yet have scientific data to quantify the impact, anecdotal observations of improved storm water filtration have been observed.

The pavement is also ingeniously designed to collect rain water from roofs, sidewalks and streets and transport it to a concrete inlet near the tree.  This inlet not only serves as a water reservoir for the tree but also proved very beneficial in terms of storm water management. 

Fig. 5: Establishment of storm water infiltration with biochar into existing tree lines. 

This latter benefit brought unanticipated support for the use of structured soils with biochar in urban environments from Sweden’s Water Department personnel.  The Water Department asked Embrén to provide an overview of their landscaping initiatives and the impact on storm water to more than 300 departmental employees.  Many workshop participants expressed shock and pleasure that the landscaping crew was able to move so quickly and effectively to implement these improvements which provided multiple intra-departmental benefits while also saving money. Given the potential impact on human health, Water Departments, in contrast, tend to have significantly more regulatory hurdles to navigate before changing any existing practices. Fortunately harvesting and filtering roof and street water for trees was a comparatively straight-forward process which not only improved the lives of trees, but reduced costs beyond just the Landscaping Department.

The Swedish National Road and Transport Research Institute (VTI), an independent and internationally prominent research institute in the transport sector, also became a supporter.  After 10 years of observing the results and impacts of the structured soils, they have given their seal of approval on this urban tree planting technique.

Word about the beneficial impact on storm water management even reached the U.S. Environmental Protection Agency (EPA) as they were gathering best practices from around the world.  The EPA invited the Stockholm team to Washington D.C. in 2014 to provide an overview on their structured soils technique.  When the team brought up their experience and plans for incorporating biochar and recycled concrete, the EPA was very interested!

Sustainably produced biochar

While all of these results have been overwhelmingly positive, there was a desire to further improve the sustainability of their landscaping efforts.  One step towards that goal was to replace the use of large stones with recycled concrete when sufficiently processed supplies can be found.  Another goal was to source biochar locally.  To date the biochar used has been purchased mostly from Germany with some also coming from Finland, England and Sweden. The biochar that is most desired is unmixed and certified by the European Biochar Certificate (EBC).

In order to be truly sustainable, however, a goal of using locally produced biochar from underutilized biomass was set. To attain this goal the City of Stockholm initiated the Stockholm Biochar Project with the objective of producing both biochar and renewable energy using urban green waste collected from municipal parks as well as city residents.  This particular biomass is often difficult to dispose of and very much underutilized.

The Stockholm Biochar Project team succeeded in winning one of the five coveted prizes of the 2014 Mayors Challenge funded by Bloomberg Philanthropies and the EUROCITIES organization which netted them 1 million euros to be used to set up their initial pilot plant. 

Once the pilot plant is up and running, estimated to be in mid-2016, the heat generated during production will be added to a local or district heating network. Biochar will be used both by city residents and by city landscapers in public parks and urban tree beds. The pilot plant will be able to produce 300 tonnes of biochar per year. At full scale biochar production will reach 1,500 tonnes per year!

Fig. 6: The first trees of Stockholm planted into biochar containing substrates.

While biochar has not yet become an acknowledged offset product in Sweden, those involved in the biochar world in Sweden are working to make this happen. Some of the big energy companies in Sweden are also very supportive.  Given that the use of biochar has significantly improved storm water quality, many involved in the water treatment industry are also backing the biochar movement in Stockholm.

When asked what advice he might give to other cities about the use of biochar in urban soils, Embran enthused: “Dare to try it and you will be convinced!”  Indeed the city has already provided inspiration and education for many landscapers looking to replicate the success they have had with long-lived, healthy urban trees: Once local biochar production is up and running, they plan to host an international conference and invite urban planners and landscapers from around the world to come and learn how to implement their own closed loop, carbon negative landscaping systems.

To read more please visit the Stockholm Biochar Project website.

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Jim Ippolito, Art Donnelly and Jim Grob

The discovery of elevated fertility of the Amazonian Terra Preta soils was widely reported in the news media starting in 2006, coinciding with the initial public awareness of the existential threat of climate change. Small groups of people worldwide seized on the idea of Terra Preta and biochar as a climate solution, and began to publicize and act on the promise of biochar. Recently, a few individuals in Seattle, WA (USA) started a group called “SeaChar” to inform the public about biochar and encourage its use. Initially, they encountered skepticism about adding what most people considered to be fuel to soil. With no existing biochar trials in the region to point to, SeaChar decided to set up a scientific field trial at a local college as a way to learn for themselves whether biochar was valuable as a soil additive. This article reports on the results of that trial and some of the challenges encountered along the way.   

Technical Abstract

Biochar land application research in elevated rainfall areas (980 mm annual rainfall) of the U.S. Pacific Northwest is lacking. A proof-of-concept field study examined the effects of spruce-pine-fir wood chip biochar (slow pyrolysis; 450-500 oC; 35 Mg ha-1), dairy manure compost (105 Mg ha-1), compost + biochar (35 and 105 Mg ha-1, respectively), and a control (no biochar or compost) on glacially altered soil chemical properties and growth characteristics of vetch (Vicia spp.) and sweet corn (Zea mays L. Golden Jubilee) over a growing season. In-season liming (5.4 Mg ha-1) occurred across all the plots to raise the soil pH for adequate crop growth. Biochar, alone or applied with compost, maintained a greater amount of soil organic C and, when combined with lime, acted more effectively than control conditions at increasing soil organic C. Biochar and compost + biochar treatments reduced Mehlich-III extractable Zn and Cu concentrations, although the concentrations were an order of magnitude greater than those considered minimal for crop growth. There was no statistical difference in vetch or corn yields among treatments. However, the compost + biochar treatment did increase vetch total N and Mg content, as well as corn Cu content, as compared to other treatments. Overall, observations suggest that co-applying biochar with an organically-rich material like compost could be beneficial without compromising environmental quality. 

The SeaChar Field Trial

Our project was initiated by the Seattle Biochar Working Group (SeaChar), founded in Seattle Washington in 2008 by two area residents (metal-work artists Art Donnelly and Don Hennick) and many volunteers (including Jim Grob, James Whittaker, Sue Dickson, and Steve Tracy). The purpose and motivation of SeaChar was to test the effects of biochar on local soils and to involve volunteers in a citizen science project that would educate students and others about soils and biochar and, if the results showed promise, use them to promote biochar for urban farming and local food production.

Left: SeaChar founders Don Hennick and Art Donnelly with a TLUD stove they designed. Right: Informational placard at the SeaChar Carbon Garden at South Seattle Community College.

In 2009, SeaChar volunteers reached out to the campus of South Seattle Community College, Seattle, Washington, and a research location was secured. Much of the outreach for working with campus administration was facilitated by SeaChar board members Vivian Scott and Anita Hornby. The approximately 0.10 ha site was located between a parking lot and a fence at the south end of campus, largely covered in blackberry bramble cleared prior to site establishment. Soil on the site was likely composed of a mixture of native soil along with fill removed during parking lot construction.

Volunteers carefully apply compost, lime and biochar to the various blocks.

Following site selection, Seachar volunteers reached out to soil science and biochar experts from both academia and the biochar technical community. The experts responded and provided guidance on plot management and treatment amounts. However, many of the experts were a long distance away and thus were not able to visit the site in person. The project utilized four treatments replicated twice in the field in a randomized complete block design (i.e. all treatments applied to specific plots, established at random, within two blocks). Blocking treatments is to account for on-site variability, dividing blocks into relatively homogenous subgroups or blocks and laying out the different treatments in a random fashion within those blocks. The variance in the data is theoretically reduced and thus the study can focus on differences between treatments instead of differences between blocks. (see http://www.biochar-international.org/sites/default/files/IBI_Biochar_Trial_Guide_final.pdf). 

Prior to site establishment, the location was an untended blackberry bramble that needed to be cleared. A one-time treatment application was utilized as suggested by Dr. Julie Major, as this approach was consistent with other research approaches worldwide. This approach would allow for comparisons to be made between our project and others. In addition to biochar, compost was used as a treatment as this had been advocated by others (e.g., Dr. Bruno Glaser from Germany) for a systems approach to organic amendment comparisons (e.g., compost vs. biochar). During the project, we were fortunate to enlist the help of Mr. Jim Grob to collect soil and plant samples on a relatively routine basis. Finally, we intended to conduct the research for a minimum of three harvests, and had a five-year lease with the College. However, we completed the work only after two field seasons and one harvest due to an in-field blocking error that was beyond our control after plot establishment (see the Challenges to Community-Based Research section, below).

Background

Social objectives

The SeaChar project had both social and scientific objectives.  One important social objective was to counteract negative perceptions of biochar that SeaChar volunteers had encountered in their efforts to promote the use of charcoal as a soil amendment.  Despite the apparent success of biochar use in tropical soils, some gardeners worried that the effects of biochar in their soils could be detrimental. SeaChar wished to show that biochar was at least not detrimental so that local gardeners could feel confident that it was worthwhile to try in their own soils.  SeaChar also wanted to involve students and volunteers in learning more about scientific approaches to soils and urban farming.  Finally, SeaChar wished to establish the test plots in a publically accessible area so that the general public could observe differences, if any, between soils amended with biochar and the alternative treatments.

Volunteers of all ages took part. Left: Elementary students mulched the paths. Right: The Seattle Youth Corps pitched in to pull weeds.

Science objectives

The scientific objectives of the SeaChar field trial were primarily to address a knowledge gap in biochar research: few, if any, studies have been performed regarding biochar effects on the younger, relatively unweathered soils of the cool, humid Pacific Northwest.  Our objective was to, via a proof of concept study, determine the effects of a one-time, realistic biochar or compost application, or a biochar and compost co-application, on soil, vetch (Vicia spp.) and sweet corn characteristics near Seattle, Washington, USA.

Increases in the soil nutrient status have been observed in relatively unweathered to highly weathered soil systems (Table 1).  Biochar application may increase available nutrients across a wide array of soil types because biochar-associated elements may be present as soluble salts (Cao and Harris, 2010; Knicker, 2007) or selectively sorbed on exchange sites (Ippolito et al., 2012a; Novak et al., 2009a; Namgay et al., 2010), both of which could provide nutrients to the soil solution and thereby to plants.  Thus, biochar may act similarly to a fertilizer for some nutrients as presented in Table 1, yet it must be kept in mind that biochar would be equal to a relatively low grade fertilizer.

Left: Observing different growth patterns. Right: Grinding corn stalks for analysis.

Table 1.  Positive changes in soil nutrient status in relatively unweathered to highly weathered soil systems following biochar application, as compared to soils that did not receive biochar.

However, increases in plant-available nutrients do not always occur.  In some instances decreases occur, while in others situations no change occurs as illustrated in Table 2.

Table 2.  Negative or neutral changes in soil nutrient status in relatively unweathered to highly weathered soil systems following biochar application, as compared to soils that did not receive biochar.

Soil and plant responses to biochar application have been related to changes in soil fertility, quantity of initial available soil N, and biochar chemical characteristics (Spokas et al., 2012).  Thus, there are a plethora of variables that determine plant response to biochar application.  More specifically, these include biochar characteristics (e.g., available nutrients, pH, ability to attract or repel water, etc.), biochar application rate and the immediate effect on soil water relations and microbiological activity, the initial soil fertility status, and plants to be grown.  The aim of the current study was to identify some of the variables affecting plant growth when utilizing biochar.  Compost was also utilized in the study since it is a readily available organic source of nutrients, used by many homeowners, and at the time of study establishment was being suggested as a comparison material to biochar.   

Materials and Methods

Biochar and Compost Characterization

The biochar, supplied by Alterna Biocarbon (Prince George, British Columbia, Canada), was created from a spruce-pine-fir wood chip feedstock mixture using slow pyrolysis at a temperature of 450-500 ^oC.  Dairy manure compost was supplied by Bailey Compost (Snohomish, WA).  Biochar and compost chemical characteristics are presented in Table 3.

Table 3. Properties and total elemental analysis of the biochar and compost, and properties and Mehlich-III elemental analysis of the soil.Analyses are described in the appendix.  ND = not determined.

Experimental Setup and Design

The experimental site was located on the South Seattle Community College Campus, Seattle, WA (lat. 43^o 32’ 41.98” N, long. 122^o 21’ 8.35” W; elevation 100 m; 980 mm annual rainfall), and was established by a group of aforementioned volunteers with the guidance of a research team.  Soils at the research site were classified by the USDA-Natural Resource Conservation Service as Inceptisols and were part of the Everett (sandy-skeletal, isotic, mesic Humic Dystrochrepts) and Alderwood (loamy-skeletal, isotic, mesic Aquic Dystrochrepts) soil series (University of California, Davis Soil Resource Laboratory, 2008).  Inceptisols are relatively young soils and are not weathered to a great extent.  Their productivity can vary widely depending on location.  Inceptisols in the US Pacific Northwest are typically quite fertile (Brady and Weil, 1999).  Background soil characteristics are presented in Table 1 above.

We chose to follow analysis similar to other biochar research projects at the time the study was established.  We evaluated changes in the soil by measuring pH, EC (electrical conductivity), nutrients and carbon content. We analyzed impacts on the plants by measuring plant growth and plant nutrient content.

In June 2009 and prior to treatment application, eight 4.5 m x 4.5 m plots were established. Four amendment treatments included the following: 1) control (no biochar or compost application), 2) composted manure applied at 105 Mg ha^-1 (dry wt.), 3) biochar applied at 35 Mg ha^-1 (dry wt.) application, and 4) composted manure + biochar co-application using rates identical to manure-only and biochar-only treatments. Figure 1 illustrates the complete block design with two replicates of all treatments within each block.  This experimental design is utilized to account for and reduce in-field variability.  In our case, there was in-field variability from block 1 to block 2.  Thus, within each block the in-field variability is reduced, and all treatments studied (in our case, four) are placed, at random, within the block.  With this experimental setup, differences between treatments can be discerned without the effect of in-field variability.

Figure 1.  Experimental plot setup using a randomized complete block design with two replicates.

Treatments were hand-applied  in July 2009, and all treatments were rototilled into the soil to a depth of 10 to 15 cm.  A one-meter, unplanted border separated all plots to help keep plots separate from one another.  In October 2009 all plots were seeded with vetch.  The vetch was grown in order to potentially improve soil N dynamics.  The vetch was mowed to a height of ~ 10 cm on May 14, 2010; clippings were collected, weighed for yield determination, and analyzed for nutrient content.  Analyses are described in the appendix.

Following vetch collection, lime was hand-applied to all plots at a rate of 5.4 Mg ha^-1 to increase the initial soil pH (5.2) to optimal for corn growth (pH 5.8 to 6.2; Hart et al., 2010).  Then, the remaining vetch and lime were incorporated to a depth of ~ 10 cm with a rototiller.  An additional soil sample was obtained on May 21 to ascertain the liming effect, and all soil analyses were performed as previously noted.

All plots were hand planted with sweet corn (Zea mays L. Golden Jubilee) on June 5.  We chose sweet corn because it grows readily in this environment and is relatively easy to cultivate and tend.  In addition, both the USDA-Agricultural Research Service and Washington State University Extension services were also using corn or  sweet corn in their biochar plot testing.  Volunteers routinely monitored the soils over the growing season by collecting samples from all plots to a depth of 0 to 30 cm.  Whole corn plants were harvested on Sept. 29 by removing plants within a 4.2 m² area within each plot.  The precise harvested area needed to be known in order to convert yield from such a small area to units known by most corn producers (e.g., megagrams of corn per hectare; Mg/ha = metric tons per hectare).  The corn was weighed and then the ears were removed, counted, and weighed.  Corn plants without ears were chopped and a subsample sent to the laboratory for nutrient analysis.  Analyses are described in the appendix. 

Statistical Analysis

Statistical analysis was performed on all soil and plant data to test whether or not differences existed between the observed averages of treatments.  In our case, we tested whether or not differences existed between the observed averages from the control, biochar, compost, or biochar+compost treatments for the measured soil and plant constituents.  For plants, we used a statistical method called analysis of variance to test whether or not differences existed between treatments.  This statistical approach is used to compare the averages (and the variability of those averages) of more than two treatments.  For the soils, we used a statistical method called “split with time” because this type of design tests whether or not differences existed between treatments, or within a given time period, or if differences existed both with treatment and time.  For instance: we may be interested in knowing whether one treatment increases a soil nutrient as compared to other treatments (exactly like how we statistically analyzed the plant data); at the same time we may want to know if time affects a soil nutrient; and we may be interested in how the treatment and time combine to affect a nutrient (e.g., one treatment could increase a soil nutrient over time while another treatment could decrease a soil nutrient over time.  This is called the interaction.  We did not observe any significant interactions in this study).       

We used a 95% confidence interval for all statistical analyses.  This means that when differences existed between treatments, we were 95% confident that they really did exist.  If significant differences were present between treatments, we then calculated a number called the Least Significant Difference (LSD; Steel and Torrie, 1980).  The Least Significant Difference number indicates what value is needed to see a significant difference between average values of treatments, again with 95% confidence. 

Results and Discussion

We analyzed the soils for organic C, total C, total N, NH₄-N, NO₃-N, pH, EC, and plant-available Ca, Cu, Fe, K, Mg, Mn, Mo, P, and Zn.  We analyzed the vetch and corn for yield and for Al, Ca, Cu, Fe, K, Mg, Mn, Mo, Na, P, S, Zn, C, and N content.  Only the significant results are presented below.

What Happened to the Soil ?

Organic Carbon
The biochar and compost + biochar treatments contained a greater percentage of soil organic C as compared to the control (Fig. 2a).  The biochar and compost used in the study contained 67.4% and 25.2% C, respectively.  Given the biochar and compost application rates of 35 and 105 Mg ha^-1, and the initial soil organic C content of 2.80%, the biochar, compost, and biochar + compost applications should have increased the soil organic content by 1.39, 1.43, and 1.82 times, respectively.  These estimated increases would equate to 3.92, 4.00, and 5.10% organic C per treatment, which are very comparable to values shown in Fig. 2a.  Similar results have been observed in other studies (see Ippolito et al., 2014b, c).  Increasing organic C content in soils is typically linked with improving the ability of soil to retain nutrients and moisture, and thus likely improving plant productivity.  If the organic soil carbon content increases continues over a longer time scale due to improvements in nutrients, water, and plant productivity, the carbon sequestration potential through biochar soil amendment could much exceed the amount of carbon sequestered in biochar alone and could thus potentially become an important climate mitigation strategy.

The effect of biochar and compost + biochar treatments in elevating soil C was maintained over the study period, which could have been due to the cool site conditions limiting microbial degradation of the organic C sources.  However, this result also supports the findings of Lentz and Ippolito (2012) that biochar may be recalcitrant and help protect compost from further microbial degradation.  Indeed, recent research (Lentz et al., 2014) suggested that biochar may reduce soil microbial processes that affect organic matter degradation, thus leading to maintenance and accumulation of C from other sources such as compost. Opposite to these findings, Wardle et al. (2008) and Hamer et al. (2004) suggested that decreases in soil organic C content may occur because some biochar C may be degraded by microorganisms or that biochar stimulated microorganisms to degrade native soil organic C.

 
Figure 2.  Changes in soil A) organic C content and B) pH over the study period in control (no added amendment), biochar (35 Mg ha^-1), composted manure (105 Mg ha^-1), and biochar + compost (35 and 105 Mg ha^-1, respectively) amended soils. 

pH
The biochar, compost, and compost + biochar treatments all had a slight liming effect, compared to the control, as the soil pH significantly increased from application date (June 2009) to the following spring (5/14/10; Fig. 2b).  This was caused by both materials having a pH of 6.8 as compared to the soil pH of 5.2 (Table 3).  Following lime application (5/14/10), the soil pH in all treatments increased as expected.  However, the compost + biochar treatment showed the greatest pH shift over time, and maintained a greater soil pH as compared to all other treatments including the control.  These observations were likely due to both biochar and compost containing some buffering capacity, or their ability to prevent the soil pH from decreasing. 

Compost has been shown to increase the pH of acidic soils (Alvarenga et al., 2008, 2009) and thus can be an effective liming agent.  Because of its neutral to basic pH, biochar may also be used as a liming agent (Kloss et al., 2012), to reduce soil acidity (Yuan and Xu, 2011; Uchimiya et al., 2012).  The pH of biochar is strongly influenced by pyrolysis temperature.  Enders and Lehmann (2012) observed an increase in pH of several biochars (cow manure, annual biomass, woody biomass) when pyrolysis temperature increased from 300 to 600^oC.  Increasing pyrolysis temperature typically removes acidic organic functional groups and causes biochar to become more basic in pH (Novak et al., 2009b; Li et al., 2002; Ahmad et al., 2012; Cantrell et al., 2012)and causes the nutrients to be in metal oxide, hydroxide, or carbonate form, resulting in elevated pH values (Cao and Harris, 2010; Knicker, 2007).

Figure 3.  Changes in plant-available (i.e. Mehlich-III extractable) soil A) Mg, B) Zn, and C) Cu over the study period in control (no added amendment), biochar (35 Mg ha^-1), composted manure (105 Mg ha^-1), and biochar + compost (35 and 105 Mg ha^-1, respectively) amended soils.

Plant-Available (i.e., Mehlich-III Extractable) Mg
Plant-available Mg increased over the course of the study (Fig. 3a), likely due to a combination of compost and lime application.  The compost and compost + biochar treatments contained greater extractable Mg as compared to the control or biochar alone treatments.  In addition, the concentration of extractable Mg increased following lime addition on May 14.  The Mg concentration could have increased due to the presence of Mg in lime, or because the soil pH increased and subsequently increased Mg availability (e.g., see Whiting et al., 2011).  Although we did not measure the Mg content in lime, in either case, increasing Mg availability could be construed as a positive in terms of plant growth.

Plant-Available (i.e., Mehlich-III Extractable) Zn and Cu
The biochar and compost + biochar treatments reduced plant-available soil Zn and Cu concentrations, likely due to biochar sorbing and making Zn and Cu partially unavailable for plants (Figs. 3b and c).  However, plant-available Zn and Cu concentrations in all treatments were at least an order of magnitude greater than concentrations considered low for crops (1.6 to 3.0 mg Zn kg^-1 soil, and < 10.0 mg Cu kg^-1 soil; Espinoza et al., 2006).  Biochar may retain nutrients via several mechanisms including entrapment of dissolved nutrients in water (Lehmann et al., 2003), surface sorption via surface groups, and electrostatic adsorption (i.e. cation exchange capacity; CEC).  In the current study, biochar CEC could develop during pyrolysis or when the product was exposed to air and water, creating oxygenated surface functional groups (Briggs et al., 2012; Chan and Xu, 2009).  Additionally, Cu and Zn may be physically trapped in internal biochar pores.

However, it was more likely that adsorption of Cu and Zn occurred on biochar.  Beesley and Marmiroli (2011) observed a significant decrease in leachate Zn content from soil (~ 300 mg L^-1) when leachate was subsequently passed through a biochar matrix (~ 10 mg L^-1).  The authors speculated that Zn was sorbed on outer surfaces of biochar, and when those sites were saturated with Zn, Zn sorbed onto inner pore surfaces.  Sorption of Cu by biochar has been researched to a slightly greater extent.  Borchard et al. (2012) suggested that oxygen-containing functional groups present in biochar are responsible for overall sorption.  The authors found that Cu interacted chemically with biochar and physical interaction (i.e., entrapment) was negligible.  Ippolito et al. (2012b) showed that, in part, Cu was bound to biochar via reactive organic functional groups on the biochar surface.  Uchimiya et al. (2012) showed that by removing these functional groups from biochar, the retention of elements such as Cu were also reduced.

Figure 4.  Changes in vetch A) total N, B) Mg, C) Zn, and corn D) Cu concentrations grown in control (no added amendment), biochar (35 Mg ha^-1), composted manure (105 Mg ha^-1), and biochar + compost (35 and 105 Mg ha^-1, respectively) amended soils.

What Happened to the Plants?

Vetch Biomass
Vetch biomass did not significantly vary across treatments.  Biomass for the control, biochar, compost, and biochar-compost treatments were 2.2, 1.9, 2.3, and 2.1 Mg ha^-1, respectively.  There was no significance difference observed across the treatments.

Vetch Total N
Co-applying compost and biochar caused a statistically positive synergistic effect in vetch total N content as compared to applying both materials separately (Fig. 4).  This observation could be related to biochar leading to more efficient compost- or soil-borne N use by vetch, or by potentially increasing microorganisms involved with N fixation as suggested by Ducey et al. (2013).  In support of this hypothesis, two growing seasons following application of biochar (22 Mg ha^-1), manure (42 Mg ha^-1), or biochar-manure at the same rates, Lentz and Ippolito (2012) observed an increase in corn silage total N uptake in biochar-manure treatments as compared to the biochar or control treatments.  Lentz et al. (2014) explained the increase in N uptake as biochar maximizing net N mineralization in the presence of manure; mineralization is the conversion of organic N sources to NH₄-N by microorganisms, with NH₄-N a form of N available for plant uptake.  

Similarly, Kammann et al. (2011)  observed greater N use efficiency (e.g., the ability of a plant to use more N that is present in the soil) in quinoa (Chenopodium quinoa Willd.) when grown in the presence of biochar (100 and 200 Mg ha^-1) and N fertilizer (100 kg ha^-1) as compared to N fertilizer only.  Chan et al. (2007) also observed a similar response in a biochar (0, 10, 50, and 100 mg ha^-1) and N fertilizer study (100 kg ha^-1).

Vetch Total Mg and Zn
The vetch grown in compost and compost + biochar treatments contained greater Mg as compared to the control (Fig. 4b), similar to the soil results.  This result may be important because improving Mg content in plants such as vetch could help reduce the incidence of grass tetany in areas like western Washington State (Hart et al., 2009).  Grass tetany is a disease in ruminant livestock such as beef and dairy, and involves a Mg deficiency leading to symptoms such as irritability, muscle twitching, staggering, collapse, coma, or death.

In addition, the vetch grown in biochar and compost + biochar contained less Zn as compared to the control (Fig. 4c).  Vetch grown in the compost treatment also contained less Zn than the control, which was unexplainable because Mehlich-III extractable Zn concentrations were similar between the control and compost treatment early in the study. 

Corn Biomass
Similar to the vetch biomass findings, corn total biomass did not significantly vary across treatments.  Biomass for the control, biochar, compost, and biochar-compost treatments were 4.8, 4.8, 4.6, and 5.1 Mg ha^-1, respectively.  The number of corn ears and corn ear weight also did not vary significantly among treatments.

Corn Cu
The compost + biochar treatment increased corn Cu concentration over the other treatments (Fig. 4d), although all treatments contained adequate Cu levels (3 mg kg^-1; Fageria, 2001).  The increase in corn Cu concentration in the compost + biochar treatment was surprising, and unexplainable, because this treatment contained the lowest concentration of Mehlich-III extractable Cu over the course of the study (Fig. 3C).  Namgay et al. (2010) showed an opposite response, whereby corn shoot Cu content was reduced by biochar application.  The authors attributed the reduction due to Cu sorption by biochar, while Ippolito et al. (2012b) proved that biochar can sorb Cu onto either the surface or precipitate Cu as a Cu-carbonate or Cu-oxide mineral within the biochar matrix.  Regardless, the compost + biochar treated corn Cu content was similar to that found by Moore et al. (2014) in a survey of 39 producer fields, while the control, biochar, and compost only treatments were lower than concentrations found by Moore et al. (2014).  This suggests that the compost + biochar treatment may be beneficial in terms of supplying Cu to the plant even when soil extractable Cu concentrations are low.

Social Impacts

The SeaChar project successfully engaged a number of students from South Seattle Community College in conducting the field trial. Community colleges serve a widely diverse group of students, which allowed SeaChar to reach a more diverse segment of the public than might have been possible at another location.  In addition to the educational impact of directly involving students in plot establishment and tending, SeaChar brought additional groups on campus to help with weeding the plots, including several groups of elementary grade and high school students.  The test plot became part of the campus Earth Day event and was a focus of interest for a Permaculture Convergence event.  SeaChar was satisfied that the test plot results were sufficient to reassure local gardeners that biochar would not harm soils and that it would be safe to experiment with biochar for potential soil improvement and soil carbon sequestration.

Challenges to Community-Based Research

The current in-field project initially began with four blocks, not two.  We had good intentions at the onset but did not, initially, properly account for site variability.  Instead of blocking with site variability, blocks were unfortunately established across site variability (i.e. vertically instead of horizontally as in Figure 1).  In Figure 1, a parking lot was adjacent to Block 2, and soils in block 2 were likely affected to a greater degree (as compared to Block 1) due to disturbance from parking lot development.  Our error in the initial blocking mistake was only noticed after data were statistically analyzed long after plot establishment. 

Another challenge of working in a community college environment is the high turnover of students who may only be on campus for one or two years. Faculty and staff have full schedules and may not find it easy to devote sufficient time to adequately supervise a long-term field trial, despite initial assurances. While the project received advice and support from biochar researchers outside the area, in retrospect, it would have been best to involve a principle investigator who was able to visit the site location before plot establishment to make a better determination of an optimal design layout, including a more thorough background soil collection and analysis prior to plot establishment. 

Based on our experiences we strongly advise that community based groups that wish to evaluate biochar recruit the assistance of a local scientist as lead investigator to guide the study.  We were fortunate enough to re-block the treatments into two blocks (due to sheer luck of how we initially set up the study) and salvage a portion of our study; others may not be as fortunate.

Summary

Biochar application, either alone or in tandem with compost, provided an increased and sustained over-the-growing-season amount of soil organic C as well as acted like a liming agent to maintain a more optimal soil pH for crop growth as compared to the control soil.  This observation suggests that the pH effect of biochar and lime co-application may be maintained for a longer period of time as compared to lime application alone, and thus may provide a long-term cost savings to producers.  Biochar and biochar + compost applications did result in lower extractable soil Zn and Cu, but nutrient concentrations remained an order of magnitude greater than minimum concentrations for crop growth. 

In this study, plant growth did not suffer due to biochar application.  In addition to these positive attributes, the compost + biochar treatment increased the total N and Mg content in vetch and improved corn Cu content over other treatments.  The improvement in vetch Mg content alone, associated with biochar-compost co-application, and may improve forage quality in areas prone to grass tetany issues.  Thus, the co-application of compost + biochar could be more beneficial than biochar alone for maintaining or improving plant and soil quality

Future Work

The research site could potentially be utilized in the future to track long-term changes in soil characteristics as well as plant responses, similar to the constituents measured in this study.  Perhaps such detailed analyses would not have to occur, and plants and soils could be monitored once yearly or bi-yearly at harvest.  In addition, it would be interesting to measure changes in the soil microbial community status, since microorganisms play a major role in N transformations, organic C degradation, and cycling of other macro- and micro-nutrients.  Additional, future on-site research could also include a re-application of all treatments as this approach is not often followed in research programs world-wide.  Future, community based studies with regards to replicated field trials or more simple demonstration experiments could very well support this approach in terms of local ownership.

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APPENDIX

Biochar and Compost Characterization Methods

The biochar and compost total C and N were determined by a laboratory dry combustion method (Nelson and Sommers, 1996), whereby C and N are converted to CO₂ and N₂ gases and analyzed using an instrument called a Thermo-Finnigan FlashEA1112 CN analyzer (CE Elantech Inc., Lakewood, NJ).  pH and electrical conductivity (or salt content) were determined on a saturated paste extract (Thomas, 1996; Rhoades, 1996), where each material is mixed with deionized H₂O to make a mixture that glistens on the surface, flows only slightly when tipped, and the mixture slides cleanly off the mixing implement.  NO₃-N and NH₄-N concentrations were determined using a 2M KCl extract (Mulvaney, 1996).  NO₃-N readily enters the extracting solution, while NH₄-N is typically held more tightly onto biochar or compost particles.  In order to remove NH₄-N, scientists typically add excess of another positively charged ion such as K, in the form of KCl.  The K replaces NH₄-N on the material, the NH₄-N enters the extracting solution, and then both NH₄-N and NO₃-N can be analyzed in the laboratory using chemical methods that develop specific colors for each constituent.  The organic N content of biochar and compost was determined as difference between total N and inorganic N (i.e. NH₄-N + NO₃-N content).  Total metal concentrations were determined using by a very strong acid digestion (i.e. perchloric-nitric-hydrofluoric-hydrochloric; Soltanpour et al., 1996).  A small amount of material is placed inside a large glass test tube and the acids are added.  Afterwards, the mixture is heated to help completely dissolve the material.  The resulting solution is then analyzed on an instrument called an inductively coupled plasma optical emission spectrometer (ICP-OES).  This instrument sprays the sample into very hot gas (up to 10,000 ^oC).  This excites all of the elements present in the sample, with the elements giving off characteristic wavelengths of light specific for each element.  The intensity of those wavelengths of light can be measured and then concentration of each element in the sample can be determined.

Soil Characterization Methods

All soils were collected within the top 30 cm of each plot.  Soils were returned to the laboratory, air-dried, ground and passed through a 2-mm sieve and then analyzed for pH, EC, total C and N, NO₃-N and NH₄-N, and organic N as described above.  Soil was also analyzed for inorganic C analysis and soil organic C was determined by difference between total and inorganic C.  Additionally, soils were analyzed for Mehlich-III extractable (i.e., a measure of plant-available nutrients in acidic soils; the extraction technique is named after the scientist that invented procedure) Ca, Cu, Fe, K, Mg, Mn, Mo, P, and Zn (Reed and Martens, 1996).  Elemental concentrations were determined using an inductively coupled plasma-optical emission spectrometer as previously described.

Vetch and Corn Characterization Methods

Vetch and corn samples were returned to the laboratory, dried in an over at 60 ^oC for 72 hours, and then ground to pass a very fine mesh screen.  Total C and N concentration of the subsample was determined as described above.  A small amount of plant sample was placed in a beaker and ashed in a furnace for 5 hours at 500^oC.  After cooling, a small amount of nitric acid was added and the samples were heated on a hot plate until all of the solid material was dissolved.  Samples were then further diluted, passed through filter paper, and analyzed for total Ca, Cu, Fe, K, Mg, Mn, Na, P, S, and Zn using ICP-OES.

Please cite as:

Ippolito J A, Donnelly A, Grob J:
Anatomy of a Field Trial: Wood-based Biochar and Compost Influences a Pacific Northwest Soil,
the Biochar Journal 2015, Arbaz, Switzerland.
ISSN 2297-1114

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