MISC

The Ham shack is the name given to the area in your home or if you are lucky enough building where you get to put your amateur radio equipment.  Mine is in the corner of the living room for everyone to see.  Unlike some older hams who don’t happen to be in graduate school my system is pretty basic.  No computers – maps on the walls – just my radio and Auto Tuner

House Shack – Icom 706 with Tuner on a Random Wire

bonus text

ACKNOWLEDGMENT

            This dissertation was enhanced by the insight and guidance of several people to whom words on this page can not express my heartfelt gratitude.

First I would like to thank my captain and research advisor Dr. Harold Collins for his on going drive to see me succeed as an individual and scientist.  Throughout the PhD process he has instilled in me the need for and helped me develop independent critical scientific thinking and research skills.  His involvement in this process was not by accident and the lives I impact in the future will be a direct result of his investments.

One scientist does not make a committee and I required the help of everyone.  My ASA President, Fran Pierce was there helping me see the big picture and giving me a working example in the art of academic networking. His understanding of the whole agriculture system both private and academic is amazing.   Steve Fransen, “Grass Guy Steve” was a constant anchor of encouragement and cookies.  His thoughtful heart, knowledge of forages, and ability to walk me through statistics were priceless.  Jeff Smith and Ann-Marie Fortuna were faithful to me throughout the process despite being in a separate location.  Their insight, teaching, and labs were instrumental in my learning process.  Their doors were always open to me when visiting the Pullman campus.

Outside of my committee several people and professors helped serve as springboards for information, questions, and editing.  Dr. Julie Tarara was the size four shoe permanently implanted throughout this process.  Dr. Tarara’s editing and organizational skills helped me get over the nearly impassable hump of writing and during this period my red ink using editor became a friend.  Dr. Joan Davenport was a teacher, friend and backboard to everything phosphorus and beyond.  Dr. Bob Stevens was always there in the nick if time to call me, “almost Dr.” and answer a question or two.  Chad Kruger, director of the Center of Sustainable Agriculture and Natural Resources at WSU is proof that young men can dream, be visionary, and follow their created purpose inside the academic setting while maintaining a biblical worldview.  I am proud to be the first science PhD graduate from Northwest University; however Chad paved the way and showed us it was possible.  It was because of Chad I met Hal, therefore I am always in your debt.   Dr. Eric Steinkamp and Dr. Bill Randolph were reminders to why science matters.  Paul Stumpf would pick up the phone on the first ring to give me countless pieces of information on the dairy industry including best management practices, markets, development and a reminder that agricultural research must connect to the grower.  Master Josh Kendall is my friend.  Dr. Stephanie Baller’s encouragement meant a great deal in the hardest times although we were both in our late 30’s we finished the PhD race. These few and all the professors, staff, and graduates students at WSU have enriched my science and life over the last 4 years.

In the lab Becky Cochran was the work horse with a mothering heart.  Her attention to detail, safety and willingness to unselfishly instruct is interwoven throughout the data in this dissertation.  It is with humble gratitude I say thank you for the cookies, birthday cakes, data analysis and talks.  In the field it was also a pleasure to have all four grown Babcock family members, plus Allen, Aaron, George, Mariah, Hannah, Monica and Noah working to help my dream come true.  Every one of them will never look at cow manure in the same way again.

One can not finish a Ph.D. at 37 without the support of family and friends.  My wife Mary may not have a PhD next two her name but it is because of her support at home mine does.  She received the brunt of my stress and frustration.  She dealt with the dropped balls on the home front.  She fixed breakfast, lunch and dinner, washed my cow manure clothes and put her hand on my shoulder when I wanted to quit.  It is because of her support this dissertation is dedicated to her and the kids.  Macuen, Joash, and Maizy were patient and as understanding as kids can be during this process.  I thank you for still loving your Dad when I was not very lovable.  My immediate family Marv and Linda Streubel, Dusty and Raeann Bruland, Dave and Esther Clark, Bunny and Wayne Winsauser sacrificed time, finances, and hunting trips to get my family through.  I can not express on paper my sincere thank you.    A thank you to the friends from King Mountain, Grandview A/G, Bethany Pres, Harvest Heights and the Northwest Network for believing in a vision.  My running group – Dr. Peck, Dr. Erin, Patrick, Tim and the Ncredibles – you helped keep me moving.    Thanks to everyone who believed in me even when I did not believe in myself.

I close with this reflection:

Dear God,

It was my personal relationship with you which led me down this academic road.  You designed me on purpose for an epic part to play in this life.  My belief in you and your guidance in my life has not been a crutch but my ongoing strength.    You have sustained me.  I acknowledge your divine role in getting me here to the finish line.  As this dream becomes a reality I await in trusting expectation for the next step to unfold.    My heart remains the same “I exist to restore hope to my generation, anyone I meet, anyway I can, anywhere I go, living dangerously and recklessly trusting Your Word, because Your love compels me to do no less.”

BIOCHAR: ITS CHARACTERIZATION AND ULITILY FOR RECOVERING PHOSPHORUS FROM ANAEROBIC DIGESTED DAIRY EFFLUENT

Abstract

By Jason Dale Streubel, Ph.D.

Washington State University

April 2011

Co-Chairs: Harold P. Collins and Francis J. Pierce

Biochar, a carbon-rich by-product of  low oxygen combustion pyrolysis used to produce alternatives to petroleum-based energy.is a  We evaluated

biochars made from four feedstocks regionally available (wood pellets [Pseudotsuga menziesii (Mirb.) Franco], softwood bark [P. menziesii], switchgrass straw [Panicum virgatum], anaerobically digested fiber) to determine their effect on five Washington soils.  Regardless of feedstock, biochars significantly raised the pH of all soil types. Biochars also increased soil C and water holding capacity (WHC) at the 20 to 40 Mg ha-1, depending on soil and feedstock type. Nitrogen mineralization rates decreased in three of the five soils across all feedstocks.  There were significant correlations (r ³0.9) between amendment rate and C increase in the amended soil regardless of feedstock or soil type.  Results demonstrate that in temperate soils, biochar feedstock may not be as important a variable as soil type for increasing C content and pH, but feedstock can influence N mineralization and WHC.

Rising production costs and environmental concerns over phosphorus fertilizer and dairy nutrient management have prompted investigations of phosphorus recovery from anaerobic digesters.  We evaluated the use of biochar produced from anaerobic digested fiber (ADF) for sequestering phosphorus from dairy lagoons and its impact on a sandy soil.  The biochar reduced P in dairy effluent by 32%.  The sequestered P was predominantly in plant-available inorganic orthophosphate.  The resulting fiber-coated biochar increased C mineralization, sodium bicarbonate (Olsen) and water-extractable P levels.  Nitrogen mineralization rates were not significantly different from the control in three of the six treatments. Results demonstrate that biochar from anaerobically digested fiber can reduce P from dairy effluent and that the P-coated biochar may be a viable source of alternative P fertilizer.  Thus, biochar produced from anaerobically digested dairy manure fiber can reduce and recover phosphorus from anaerobically digested effluent lagoons and increase plant available phosphorus in the soil, which has potential environmental benefits.  This work provides new knowledge and understanding of biochars from Northwest feedstocks and their ability to recover P.

TABLE OF CONTENTS[MSOffice1]

ACKNOWLEDGMENTS………….…………………………………………………………… iii

ABSTRACT……………………………………………………………………………………. vi

LIST OF TABLES ……….………………………………………………………………………vi

LIST OF FIGURES……………….…………………………………………………………….. xi

ABBREVIATIONS…………………………………………………………………

DEDICATION………………………………………………………………………….………. xii

CHAPTER ONE. GENERAL LITERATURE REVIEW………………………………………… 1

Literature Cited…..……………..………………………………………………………..26

Figures……………………………………………………………………………………..34

CHAPTER TWO. INFLUENCE OF BIOCHAR ON SOIL pH, WATER HOLDING CAPACITY, NITROGEN AND CARBON DYNAMICS

Abstract…………………………………………………………..……………………. 35

Introduction……………………………………………………………………………….36

Materials and Methods.………………………………………………………………… 38

Discussion and  Results……………………………………….………………………. 42

Literature Cited.………………………….……………………………………………… 53

Tables……………………………………………………………………………………. 59

Figures……………………………………………………………………………………. 63

CHAPTER THREE. THE ABILITY OF ANAEROBIC DIGESTED FIBER BIOCHAR TO REDUCE PHOSPHORUS IN DAIRY LAGOONS AND ITS CHARACTERIZATION

Abstract…………………………………………………………………………………. 72

Introduction………………………………………………………….…………………..

Materials and Methods.……………………………………………….…………………

Results and Discussion …………………………………………………………………

Literature Cited…………………………….…………………………………………….

Tables……………………………………………………………………………………

Appendix ……………………………………………………………………………….

CHAPTER FOUR. Phosphorous, N and C availability from biochar

amended with dairy effluent

Abstract………………………………………………………………………………….

Introduction…………………………………………………………………….………..

Materials and Methods.……………………………………………………….…………

Results and Discussion……………………………………………………….………….

Conclusions………………………………….…………………………………………..

Literature Cited………………………….……………………………………………….

Tables……………………………………………………………………………………

Figure……………………………………………………………………………….……

CHAPTER FIVE. GENERAL CONCLUSIONS…………  ……………………………………

LIST OF TABLES

CHAPTER TWO

Table 1. Selected characteristics of the five soil types used in the laboratory incubations………59

Table 2. Change in C and N concentrations of feedstocks following pyrolysis at 500oC…….…60

Table 3. Carbon and N concentrations, and changes in concentrations after acid hydrolysis, for four biochars pyrolized at 500oC. ……………………………………………………………….61

Table 4. Elemental composition and pH of four biochars and a commercial activated charcoal control used in laboratory incubations……………………………………………………………61

Table 5.Water holding capacity (WHC) of five soils at 0.1 MPa after variable rate amendment with four biochars pyrolized at 500oC……………………………………………………………62

Supplement Table 1.  Total C, C-mineralization and estimated pool sizes of the active (Ca), slow (Cs) and resistant (Cr) C pools for Quincy sand soil amended with four rates of biochar pyrolized from each of four feedstocks……………………………………………………………………..67

Supplemental Table 2.  Total C, C-mineralization and pool sizes of the active (Ca), slow (Cs) and resistant (Cr) C pools for Naff silt loam soil amended with four rates of biochar pyrolized from each of four feedstocks……………………………………………………………………………68

Supplemental Table 3.  Total C, C-mineralization and pool sizes of the active (Ca), slow (Cs) and resistant (Cr) C pools for Palouse silt loam soil amended with four rates of biochar pyrolized from each of four feedstocks…………………………………………………………………….69

Supplemental Table 4.  Total C, C-mineralization and pool sizes of the active (Ca), slow (Cs) and resistant (Cr) C pools for Thatuna silt loam soil amended with four rates of biochar pyrolized from each of four feedstocks…………………………………………………………………….70

Supplemental Table 5. Total C, C-mineralization and pool sizes of the active (Ca), slow (Cs) and resistant (Cr) C pools for Hale silt loam soil amended with four rates of biochar pyrolized from each of four feedstocks………………………………………………………………………….71

CHAPTER THREE

Table 1. Selected characteristics of the four stages in the material process.……… 42

Table 2. Concentrations of phosphorus and fiber in each experimental lagoon.

Table 3. Elemental composition of manure pellet before and after pyrolysis and three periods during the 15d filtration with effluent.

Table 4. Elemental composition of liquid and solid fraction after extractions of the initial biochar and biochar after 15d of filtration through anaerobic digested effluent used in 31P NMR analyses.

CHAPTER FOUR

Table 1. Selected characteristics of Quincy sand used in the laboratory incubations.

Table 2. Elemental composition of the untreated and dairy effluent amended biochars.

Table 3.  Total cumulative C-mineralization for Quincy sand soil with untreated and dairy effluent amended biochars.

Table 4. Statistical analysis of Olsen P for 21 day soil incubations with untreated and dairy effluent amended biochars.

Table 5. Statistical analysis of water extractable P for 21 day soil incubations with untreated and dairy effluent amended biochars.

TABLES OF FIGURES

CHAPETR ONE

Fig 1. The Phosphorus Cycle ……………………………………………………………34

CHAPTER TWO

Figure 1. Influence of biochar amendment rate on soil pH for five soil types and four biochars.63

Figure 2. Total N-mineralization at 49d of laboratory incubation as a function of soil type and biochar.  …………………………………………………………………………………………64

Figure 3. Relationships between the recovery of biochar C (a) and acid hydrolyzed biochar C (b) by the biochar C initially applied to each soil……………………………………………………65

Figure 4. Rates of C mineralization of soils amended with switchgrass biochar for Quincy sand, and Naff, Palouse, Thatuna, and Hale silt loams measured during a 225-d laboratory incubation. ……………………………………………………………………………………………………66

CHAPTER THREE

Figure 1. Barrel retort temperature recorded over the 4 hour pyrolysis period.

Figure 2. Diagram of anaerobic digested effluent biochar recovery experimental set up.

Figure 3. Change in effluent pH following the application of biochar during

a 90 h laboratory experiment. The ratio of AD effluent to biochar was 5:1.

Figure 4. Change in lagoon effluent pH after continuous flow filtration through

biochar over a 15d period.  The ratio of AD effluent to biochar was 25:1.

Figure 5. Scanning electron micrographs of biochar prior to filtration (A) and (B) biochar 15 d after used as a filter media for anaerobic digested dairy effluent.  Micrographs taken at the Environmental Molecular Sciences Laboratory, Richland, WA.

Figure 6. Change in NaCHO3 (Olsen) and water extractable P  (WXP) from biochar after continuous filtration of anaerobic digested effluent over a 15d period.

Figure 7. Liquid State NMR for biochar prior to filtration (A) and (B) biochar 15 d after used as a filter media for anaerobic digested dairy effluent.

Figure 8. 31P Solid State NMR spectra for biochar prior to filtration (A) and (B) biochar 15 d after used as a filter media for anaerobic digested dairy effluent.

APPENDIX

Appendix Figure 1 Batch Retort Pyrolysis unit constructed for biochar production

Appendix Figure 2 378 L Experimental lagoon set up in Prosser, WA

CHAPTER FOUR

Figure 1 Total NO3-N Mineralization after Day 49 in Quincy Sand

Figure 2 Total rate of CO2-C for APF Biochar and APF Biochar with Coatings

Three Rates of Application in Quincy Sand Over 217 Days Letters represent significance at >0.05

Figure 3. Sodium Bicarbonate extractable P from three amendment rates of untreated and dairy amended biochars.

Figure 4 Water extractable P from three amendment rates of untreated and dairy amended biochars.

ABBREVIATIONS

BA – Amended Biochar

BI – non-amended Biochar

BI – Initial non-amended biochar
DEDICTATION

This dissertation is dedicated to my wife Mary, sons Macuen and Joash, and daughter Maizy for they have stayed with me through this process.  I could not have survived without them.

CHAPTER ONE

GENERAL LITATURE review

Introduction

World phosphorus (P) supplies are annually mined as rock phosphate from three areas (US, China, and Morocco) and exported globally to meet fertilizer demands.  Prior to the mid 1800’s phosphorus demands were met by local resources such as animal manure, human waste, bone, or native soil reserves.  As world populations grew and food production increased additional sources of P became necessary to find and the global market for P fertilizer lead to the mining and export of bat guano (Stewart et al., 2005).   Guano exports from the Pacific Islands increased for almost 80 years until the supply was no longer sufficient or profitability.  It was at this point in the 1880’s when rock phosphate became the standard for the P fertilizer market.  As the demand for P is projected to increase over the next 50 years in response to world population growth, agricultural analysts predict the world supply of rock phosphate will be depleted within the same time period (Cordell et al, 2009).  As the supply can no longer meet demand many nations will potentially restrict rock phosphate exports.  As world supply and markets tighten our focus should be on improving the efficiencies of plant uptake, agronomics, characterization, and improving P recovery from overloaded systems (Leikam and Achorn, 2005).

P is a macronutrient whose chemistry, reactions, and cycling in the soil and surrounding environment is complex.  The fate of P in soil is a function of pH, temperature, structure, and clay content. Yuji and Sparks (2007) have presented the major theories of P reaction dynamics in soil and describe the chemistry of P, absorptions, surface complexation, and the effect of reaction time in the soil matrix on P availability.  For example they state “P absorption in soils increases with the decrease of pH” (p. 138), “the retention of P on soils is highly dependent on the physiochemical properties of the soil, for example crystalline and amorphous and aluminum oxides and organic matter (OM), clay and calcium content”(p. 139). Barrow (1974) had earlier suggested that over time P was not permanently fixed to the soil but depending on soil type could become plant available. Barrow (1974) further determined that there were significant changes in available phosphate with changes in soil temperature.

Phosphorus is either readily available to the plant via the soil solution, or fixed to the soil matrix by chemical reactions indefinitely or mineralized by soil organisms for use by plants over time (Bittman et al., 2004; Havlin et al., 2005).  Inorganic P can be present in soils as HPO4-2 (pH 5-6), H2PO4(pH 6-8) and in apatite’s such as Ca (pH < 8) and Al (pH <4) phosphate P is primarily made available to plants through the mineralization of organic material (biomass, soil organic matter, plant residue or agricultural by-products) and the dissolution of primary and secondary P minerals.  Some researchers categorize P into labile and non-labile forms where labile are used to describe the readily plant available forms and non-labile as the forms needing to undergo a breakdown in to the soil solution before becoming plant available (Sharpley and Tunney, 2000; Pierzynski et al., 2005).

The P cycle (Figure 1) includes the components of plant uptake, dissolution and precipitation as well as inputes and outputs into the soil and losses to the environment.  There are three major pathways from which P can leave the soil system. These include surface runoff, subsurface lateral flow, and leaching into the groundwater.  The effect of these pathways on the environment is increased when there are high P levels (>30 mg/L Olsen P) and excess water.  If P laden runoff enters into water ways eutrophication may result (Sharpley et al., 2005).

Eutrophication is generally described as an increase in algae growth brought on by the addition of nutrients into an aquatic system and the production of hypoxic conditions as the micro and macro biota grow and consume oxygen (Sharpley et al., 2003).  As the oxygen is consumed anaerobic zones are created.  These low oxygen zones become problematic for fish and other aquatic life.  The interactions between nitrogen and P, temperature, available sunlight and individual waterway characteristics determine the severity of eutrophication, however reduction of nutrient inputs is the best approach to control of the problem (Pierzynski et al., 2005; Sharply et al., 2000; Wendt and Corey, 1980).

Concentrated animal feeding operations (CAFO) and long-term dairying increase and often exceed nutrient loading rates particularly P (Sharpley et al., 2003).  The annual estimate of solid manure produced in the United States by swine (Sus domesticus), chicken (Gallus gallus domesticus) and dairy cattle (Bos indicus ).is 71 million Mg including 2.3 million Mg of P which is greater than the annual amount added by commercial sources (Toor et al.,, 2006). A dairy with 4000 Holsteins on site can potentially produce 110 million liters of liquid manure a year that must be disposed of in some fashion. The common practice has been application of lagoon water to adjacent agricultural fields (Hart et al., 1997). Long-term applications of manures leads to  excess P runoff and potential environmental damage by eutrophication.

One method for waste management and reduction in the dairy system is the use of anaerobic digesters (AD). Modern AD systems receive incoming manure waste in solid scrapings or liquid form and cycle the waste through an environment optimal for methane production.  The warm (25oC) low oxygen environment is idea for methogenic bacteria.  Methanogenesis, the creation of methane by microbes, is a form of anaerobic respiration(Thauer, 1998). Methanogenic bacteria do not use oxygen as an electron acceptor for respiration; the terminal electron acceptor in methanogenesis is carbon.  The methane can then be used as a biogas for energy production.  The benefits of AD to a dairy include reduced odor, and reduced greenhouse gas emissions (Kaparaju and Rintala, 2010; Hobson and Feilden, 1982; Rico et al., 2007; Gungor and Karthikeyan, 2008; MacConnell and Collins, 2009). These are on top of the economic benefits from the possible energy production sold to local energy companies.  Although AD has been researched for years there are still several issues to overcome.  In addition to production costs, permits andnutrient management plans are important issue. The waste stream is high in nutrients.  The high nutrient effluent then becomes part of the overall nutrient management practice of the dairy. The incorporation of nutrient recovery systems along side AD is a logical need.  Nutrient recovery systems can become a win-win situation for both the environment and the dairy.  If we look at non-renewable P supply its recovery from effluent exiting the anaerobic digesters, fits the scenario[MSOffice2] .  Phosphorus recovery has been a focus of research within the dairy industry for many years but applicable research results have not been forthcoming.

Knowledge of P recovery has been advanced over the last decade by the wastewater industry (de-Bashan and Bashan, 2004).  The concern of eutrophication and the need for stricter regulations has driven local state and federal agencies as well as municipalities, and private industry, to invest in research concerning P recovery.  The technologies and methods of P recovery come in different forms but all fit within three categories.  The first utilizes precipitation of P with metals.  The use of iron, aluminum, magnesium and calcite products from other industrial wastes such as a slag, or oxides from steel production have been added to wastewater or sludge that complexes with the metals by forming aggregates from precipitation allowing the resulting solid to be collected (Thomas, 1972; Matsumiya et al, 2000; Banu et al, 2008).  These metal ions have been used in various environments depending on the pH of the wastewater with aluminum utilized in systems with at  pH 3.6, and calcite used in high pH systems (pH 9.0) (de-Bashan and Bashan, 2004).  The resulting precipitate may be used for a secondary P product but may also be rendered useless for soil application because of high metal toxicity (Delhaize and Ryan, 1995).  Other approaches include the biological removal of nutrients by cultivation of microorganisms in wastewater systems (de-Bashan and Bashan, 2004). These approaches have used selected bacteria and microalgae but the costs has restricted this technology from wide spread industry use outside of the research and development arena.  The construction of wetlands to naturally filter waste waters has also been used.  This concept mimics nature for natural filtration of nutrients and contaminates (Uusi-Kamppa et al., 2000; Bashan and Bashan, 2004).  Soil characteristics, plant communities, environmental condition and flow rates are all important in the form and amount of P recovered or removed by this approach (Nairn and Mitsch, 2000;Leader et al., 2005).  One emerging technology is Enhanced Biological P Recovery (EBPR) where biologically enhanced wetlands are used for the removal of P.  As the focus of P removal technology turns towards recycling, the ability of the resulting co-products to be used as a fertilizer supplement will be a driving force. (de-Bashan and Bashan, 2004)

In response to the need for P removal in the dairy system there has been  a revitalization in the recovery of struvite (magnesium ammonium phosphate hexahydrate, NH4MgPO4-6H2O) for recovering P from various waste streams (Massey et al., 2007; Westerman et al, 2009).  Based upon the discovery of struvite formations in 13th century restrooms Schuiling and Andrade (1999) developed an approach to successfully recovery struvite from calf and chicken manure.  Australian scientists redced P concentrations by 61 mg/L from municipal waste water treatment sludge using magnesium hydroxide to force the precipitation (Munch and Barr, 2001).  The implementation of engineered cone-shaped, fluidized bed systems allowed for greater control over the struvite formation process and proved successful in swine and dairy systems (Bowers et al., 2007; Bowers and Westerman, 2005). However, the production levels were lower in dairy systems because of the higher Ca content found in dairy manure (Bowers et al., 2007). This high concentration of Ca in dairy manure and effluent complicates struvite production therefore raising cost; which in turn makes this technology not yet uniformly accepted.  Struvite formation occurs within time frames of hours to weeks depending on the type of waste stream, P concentration, temperature and concentration of Mg and Ca.

As an alterative to struvite production in dairy manures researchers at Washington State University are using polymer flocculation (Chen et al., 2009).  Flocculation is the process where small particles or suspended solids are clumped together so they fall out of suspension. Flocculation is driven by the electro-static charge of the colloids in water.  In nature, particles of negative charge repel one another preventing flocculation from occurring and keeping the particles in suspension.   Flocculation  reduces the natural charge of particles  allowing them to group together.  As the groups become larger they settle out.  The ability or inability for a particle or colloid to flocculate is determined by whether van der Waal attractions are greater or equal to the repulsive forces without extra energy input (Zhaoet al., 2009, Basu and MacDonald, 2007, Ravina 1993)

Chemical polymers can be added to solution to create the flocculation for P removal, however the process is expensive. Polymers are long branched chains which can trap particles in water.  The polymers can also be custom engineered with a chemical charge to negate repulsive forces.  The optimization of the polymer –flocculation process is research intensive as liquid and chemical chemistry are unique in any given situation (Zhaoet al., 2009; Basu and MacDonald, 2007, de-Bashan and Bashan, 2004; Revina 1993).  This technology due to its expense is still currently out of the reach of the dairy industry for on farm implementation.

PHOSPHORUS ANALYSIS

Determining the best method of P analysis within each pool adds to the already complicated issue of P solubility and availability (Yuji et al., 2005; Vadas et al., 2007; Yuji and Sparks, 2007; Cluster et al., 1972; Javid and Rowell, 2002). Each factor such as; soil, temperature, and pH contributes to P solubility and its determination.  Based on variability of environmental conditions conclusions can only be addressed at the local or regional level.  Since P is being tracked as a source of  pollution on dairies and adjacent agricultural land where manure is applied, it is important we understand the sources and sinks of P.  When attempting to account for the P in the soil, it is vital to use appropriate testing methodology.   A wrong methodology could lead to misinformation on the P pools thus negatively impacting environmental regulation and experimentation (Fuhrman et al., 2005; Sims, 2005; Rowell, 2003; Barrow, 1974).

Toor et al., (2006)reviewed  current advances in P characterization  of organic P materials including agricultural and municipal waste.  The review described the basic methods being used today for P characterization included: total P, water extractable P (WXP), physiochemical fractionation, sequential P fractionation, enzyme hydrolysis, nuclear magnetic resonance spectroscopy (NMR) and X-ray absorption near edge structure spectroscopy (XANES).  The authors provided a thorough description and explanation for each of the above methods and we refer the reader to this reference for detailed information.  The literature for soils testing of P lists three primary approaches: (NRCS, 2008; Yuji and Sparks, 2007) i) available soil P which is considered available to the plant immediately. Methods for testing available P include Water Extractable P (WXP) and Ion-Exchange Membranes (AER) (Davenport and Schiffhaer, 2007) ii) labile P (LP) which could be available to the plant throughout the growing season. Methods for testinginclude Olsen, Bray, and Mehlich 1 or 3 (Kuo, 1996). iii) non labile P (NLP) which is contained in the clays or secondary minerals which is only available to the plant as it is broken down over time, usually a very slow process.  Methods include Sequential P Fractionation, Nuclear Magnetic Resonance Spectroscopy and X-Ray Absorption Near Edge Structure Spectroscopy.

Researchers have begun to understand the need for site specific identification of P forms using the techniques available today.  Fuleky and Tolner (2006) used Olsen P to fit an absorption isotherm model for the determination of soil phosphate content.  They used non linear regression to determine the P concentration with known amounts of phosphates in individual soils.  A study on an Argentine Mollisol used Bray 1 for measuring soil available P after fertilization taking into account the individual characteristics of the soil’s  pH, particle size, and organic C and N content (Rubio et al., 2008). Turner et al. (2003) analyzed Southern Delaware sandy soils for phosphate reactivity using an electron microprobe, ZANES, and inorganic and organic fractionation methods.  They determined the individual concentrations of the P needed at a local level for developing best nutrient management plans.  As the methods have become more advanced and efficient the need for site specific testing is becoming more crucial to the agricultural community for P management.

Phosphorus concentrations are commonly using two analysis, the water extractable P (WXP) and anion exchange resins (AER) method.  WXP, also referred to as a saturated paste extraction, is a simple method to determine the amount of available P at any point in time.  This method gives a relatively accurate measurement however there is a belief that the method is subject to error (Fuhrman et al., 2005).  Since the protocol does not specify the amount of soil or extractant used results come in a wide range of variation depending upon the technical expertise of the person running the analysis. It is this subjectivity that discourages it as a preferred method. This concern was addressed by Fuhrman et al., (2005) who determined that standardizing time and soil to extractant ratios were important.  They found a significant correlation between accuracy and precision in the WXP method when a standardized procedure of predetermined amounts of water, soil, and shaking time was followed.

The use of ion-exchange membranes to monitor the immediately available P pools is a method gaining momentum throughout the research community because it gives a cumulative measurement over a one to two week period.  The membranes are placed in the soil at the desired depth then collected and taken through a standard extraction methodology yielding a relative phosphorous index.  The advantage of this method allows for more than just a point in time but a process for gathering an entire season if desired. This method has the draw back of expense. Davenport and Schiffhaer, (2007) used ion-exchange membranes in a cranberry production study and found the membranes to be a promising useable tool because of the ability to collect the plant available P over a period of time.  Early, Rabeak et al., (1999) found the method very useful when run in conjunction with P-31 NMR when determining the P levels in various soil particle size fractions. With the use of the water extractable P and resin membrane extraction methods site specific data for the immediate plant available P pool can be easily determined.

Three chemical extraction methods have been developed for determining the labile P pool which is mineralized over the growing season and becomes available to the plant. The three are Bray-1, Mehlich-3 and Olsen and their use is determined by the site specific characteristics including pH and soil mineral content.  The Bray-1 test is the most common and is designed for soils with a pH of less then 7.4. The Mehlech-3 method can be used within the same pH levels and has the additional advantage of determining macro-and-micronutrients on the same sample.  The Olsen method is best suited for alkaline, calcareous soils with a pH above 7.2.  These methods use a variety of chemicals to extract P, ranging from acetate, diluted HCL and sodium bicarbonate.  These three methods are the standard for agricultural soils.  The Oregon State Extension soil testing guide recommends Bray-1 for west of the Cascade Mountain range and Olsen east of the Cascades. They also recommend results be used only as an index and not for calculating kg of phosphate per hectare (Marx et al., 1999; Elrashidi, 2008).

The final pool for accessing P availability is the non-labile pool primary and secondary minerals which are fixed to clay particles of the soil (Figure 1).  The labile P concentrations maybe easy to test; however the ability to determine inorganic P in soil becomes costly requiring specialized equipment making determinations for research purpose only.  The common practice is to determine total P, labile P (Bray-1, Mehlech-3 or Olsen) and the  non-labile fraction by the difference between the two methods. This method provides an estimate of P in the system.

The use of P-31 Nuclear Magnetic Resonance Spectroscopy (NMR) is quickly becoming a methodology of choice for accounting for both the labile and non-labile P fractions in soils. Veeman (1997) provides an excellent review of the principals and applications of NMR.  The beginning of NMR technology and its expansion took place between 1938 and 1952 as part of the race to prefect radar (Pierzynski et al., 2005).  P-31 NMR can identify organic P in DNA and ATP found in microbial communities and micro fauna as well as the inorganic compounds found on clay particles of soil.  NMR spectroscopy depends on the behavior of the nuclei of the chosen element and how it responds when hit with a magnetic field and electromagnetic frequency.  The response is unique to its spin quantum number which comes from the number of unpaired protons and neutrons (Pierzynski et al., 2005).  Nuclei with a quantum number above 0 are viable for NMR analysis, for 31P the quantum spin number is one half. When the nuclei are under normal atmospheric conditions the nuclei are in equal states and spin randomly in a magnetic field the nuclei spin in an organized fashion in direct orientation to the magnetic field.  When aligned nuclei are bombarded with electromagnetic radiation at specific frequencies energy is absorbed by the nucleus.  The NMR instrumentation has the ability to detect the energy absorbency and determine a matching resonance.  The amount of energy necessary and radiation duration to achieve resonance depends on the strength of the magnet and the element being studied (Piezynski et al 2005).    Samples can be run in solid or liquid forms. The liquid form has better resolution of P using NMR.  The solution phase significantly enhances the clarity of the peaks making for better results (Veeman, 1997).

P-31 NMR has been used to characterize P from grassland soil leachate (Toor et al., 2003) in swine, poultry and beef cattle manure (Turner et al., 2004), soils amended with manure (Hansen et al., 2004) and to measure bioavailability and determine composition of organic P in semi-arid arable soils in the western United States (Turner et al., 2003).  He et al., (2008) used NMR to study P pool accumulation in long term manured agricultural soils.

PYROLYSIS and BIOCHAR  PRODUCTION

The thermo-chemical conversion (pyrolysis) of plant biomass is being studied in the U.S. for its potential to produce second generation liquid biofuels and the co-product biochar, which can be used as a soil amendment. Pyrolysis of biomass yields 60-70% of the original biomass as a bio-oil and 15-25% as biochar with the remaining 5-15% as ash (Graham et al., 1984; Czernik et al., 1994; Bridgwater et al., 1999; Mohan et al., 2006; Huber and Corma, 2007). Biochar is the carbon-rich product remaining after biomass has been heated at low temperatures (ca. 350°C to 600°C) in an environment with little or no oxygen (Amonette and Joseph, 2009). Biochar is currently the accepted term for pyrolysis-derived charcoal when used as a soil amendment (Sohi et al., 2010); however char, charcoal, and black carbon are also been used.

Biochar characteristics are highly variable depending on feedstock, temperature, duration of the process and pyrolysis method.  The desired products of pyrolysis are  bio-oil or syngas. The desired product  ultimately determines the process and method of production.  The basic methods of production come in four categories 1) gasification 2) fast pyrolysis, 3) flash pyrolysis, and 4) slow pyrolysis. Gasification is the burning of biomass at temperatures above 800oC in the presence of oxygen.  The primary product from gasification is syngas for energy production with little biochar yield.  Fast pyrolysis is a continuous flow system where a ground biomass is heated to temperature in less than a second and maximizes bio-oil production.  As the material is heated the resulting vapors are cooled.  In contrast to the previous technological processes, slow pyrolysis is a batch system favoring biochar and syngas production. Flash pyrolysis prioritizes the production of biochar under higher pressures with some bio-oil production as well.  This generally takes place in a batch reactor rather than under continuous flow such as fast pyrolysis. Slow pyrolysis is the traditional method used for creating charcoal.  Slow pyrolysis can be conducted in kilns, mounded soils, barrels or large pits.  Which of these processes used is determined by the desired product of the pyrolysis (Sohi et al., 2010, Laird et al, 2009).

Biochar yield as well as, the chemical and physical characteristics of biochar depends on the nature of the feedstocks used (woody vs. herbaceous) and operating conditions and environment of the pyrolysis unit (low vs. high temperature, residence time; slow vs. fast pyrolysis, heating rate and feedstock preparation, etc.).  The wide range of process parameters leads to the formation of biochar products that vary considerably in their elemental and ash composition, density, pore size, surface chemical properties, water and ion adsorption and release, surface area, pH and uniformity of biochars’ physical structure (Baldock and Smernik, 2002; Antal and Grǿnli, 2003; Downie et al., 2009; Amonette and Joseph; 2009; Krull et al., 2009; Chan and Xu, 2009).

The biochar yield from the pyrolysis of biomass is influenced by the pyrolysis temperature, its lignin, cellulose, and hemicellulose content and to a lesser extent the extractive concentrations of the feedstock (McKay and Roberts, 1982; Antal and Grǿnli, 2003). Woody biomass with high lignin contents typically produce greater char yields than those derived from herbaceous feedstocks (Demirbas, A. 2006). A common feature among the various pyrolysis processes is that C content shows a consistent increase with increasing temperature (Antal and Grǿnli, 2003; Schnitzer et al., 2007; Zabaniotou et al., 2008; Joseph et al., 2009). Biochars with large amounts of poly-condensed aromatic structures are obtained by pyrolyzing feedstocks at temperatures between 400 and 600oC (Antal and Grǿnli, 2003; Amonette and Joseph, 2009).  Temperatures above 500oC commonly produce chars with C contents greater than 80%, at temperatures between 400 and 500oC biochars have C contents that range from 60 to 80%, where biochars produced at low temperatures (< 350oC) have C content that vary from 15 to 60% (Joseph et al., 2009). For a detailed description of the thermal degradation process (dehydration, pyrolysis, graphene formation and carbonization) that transforms biomass see reviews by Amonette and Joseph (2009) and Brown (2009).

Previous studies of bio-char production have shown that chars made from herbaceous feedstocks switchgrass (Panicum virgatum L.), digester fiber, peanut hulls (Arachis hypogaea) had lower carbon contents, higher nitrogen contents, and higher pH than chars made from woody feedstocks (Novak et al., 2009; Granatstein et al., 2009).  The pH of biochar is dependent on the feedstock and pyrolysis temperature and is attributed to the chemical cracking of hemicellulose and cellulose during pyrolsis. Between 300 to 600oC organic acids and phenolic substances are created and alkali salts are formed that raise the pH of the biochar (Abe et al., 1998; Shinogi and Kanri, 2003). The higher pH of herbaceous biochars give them a greater liming impact per ton of biochar added to soil, increasing soil pH 0.5 – 1.0 pH units depending on soil type.  Biochars produced from poultry litter have a pH range of 8.5 – 10.3 that is related to the concentration of Ca and Mg as higher Ca concentrations can serve as buffers(Chan et al., 2008; Gaskin et al., 2008; Sistani and Novak, 2006).

Elemental analyses of many biochars are available in the literature; however, many of these data sets are incomplete, providing information for only a few elements which makes comparisons among biochars difficult (Baldock and Smernik, 2002; Glaser et al., 2002; Antal and Grǿnli, 2003; Gaskin et al., 2008; Novak et al., 2009; Chan and Xu, 2009). Much of the mineral content of the feedstock remains in the resulting biochar, where it is concentrated due to the loss of C, H and O during pyrolysis (Amonette and Joseph, 2009). Biochars produced from biosolids and manures are typically high in ash, for example, chicken-litter biochars can have greater than 45% mineral matter (Lima and Marshall, 2005).

There is considerable variation in the content of many elements especially N and P due to feedstock characteristics and range of production temperatures. Feedstocks typically high in N, P, K and S are sewage sludge, animal manures and biosolids.  Elemental content of the biochar produced reflects the concentration of elements in the original feedstock. Herbaceous biomass and biochars derived from biosolids and manures have N, P, K and S contents greater than woody feedstocks. Most of the N and S are lost during pyrolysis as temperatures increase from 350 to 600oC.  Significant N losses (60 – 80%) have been reported for biochars derived from sugar cane bagasse (Bagassa guianensis), rice (Oryza sativa) husks, sewage sludge, and cattle manure (Shinogi and Kanri, 2003). The availability of the remaining N and P contained in biochar is considered limited.  Bagreev et al. (2001) suggested that as the pyrolysis temperatures increase, N forms pyridine-like complexes that reduced availability.  Cao and Harris (2010) suggested that the decrease in N concentration can be attributed to volatilization during heating and that some of the N-containing structures in the biochar (e.g., amino acids, amines, amino sugars) are condensed into recalcitrant forms. The nitrogen contained in biochar does not appear to be available in plant soil matrix. Increasing rates of char amendments have led to reductions in soil nitrate release compared to un-amended soils, perhaps due to ammonium (NH4+) adsorption and the inhibiting of ammonification due to the biochar. Lehmann, et al, (2003) found that (NH4+) was adsorbed by charcoal with an increase in N uptake by rice. This effect suggests improved N conservation in soils and less off-site movement of nitrate due to leaching, as well as, a potential reduction in losses due to N2O production.

Activated carbon (AC) from various sources has been used for P removal.  Tamarind (Tamarindus indica) nut shell AC was found to successfully remove 95% of the P from a standard solution through adsorption (Bhargava and Sheldarkar, 1993). Powdered AC from hardwood was used to remove from industrial waste water effluent (Meidl, J.A, 1997).  Along with almond(Prunus dulcis) and pecan (Carya illinoinensis) shells being used as AC sources for P removal (Toles et al., 1997) poultry manure derived AC has been used to remove mercury from coal fire plants (Klasson et al., 2010;Yaji et al., 2008).  The use of AC from multiple sources from waste streams has been a proven method for removal containments, however activation of the charcoal remains expensive.  The expense has prevented the use of AC  in waste water management programs.

The adsorption of phosphate ions will depend upon the concentration and accessibility of cations found in the ash. Effects of metal ion concentrations in fly ash have been the subject of significant research (Lu, et al., 2009; Agyei et al., 2000; 2002; Namasivayam and Sangeetha 2004; Oguz, 2005; Zhang, 2007; Xue et al., 2009). The use of fly ash has been considered for its potential to remove phosphate compounds from waste water. The application of this information to biochars for the removal of relatively low concentration phosphates is of interest, because of the desire to develop an adsorbent capable of removing primarily nitrogen and phosphorous but other compounds, as necessary, from waste streams. The addition of appropriate metal ions to the structure of the biomass should aid in creating additional basic sites on the char surface which will become positively charged in solution and attract anions to the surface.

Unlike the oxidation reactions which form chemically bonded surface groups on the carbon matrix, the metallic cations may need to be contained within the carbon matrix or chelated on the surface of the carbon structure to remain stable and not dissociate into solution treated with carbon. Dissociation of specific metals can cause the loss of potentially marketable nutrients such as phosphates by the formation of insoluble salts with these metal ions, such as the interaction of apatites with calcium ions (Lu et al., 2009).

Considerable effort has been made towards understanding the primary mechanism of phosphate removal. Lu (2009) showed that while the majority of phosphate removal appeared to be due to precipitation reactions, the data also suggested that a reasonable level of adsorption can be achieved on the ash. Not much attention has been given to the simultaneous removal of ammonium and phosphate; although the work by Zhang (2007) shows significant potential in this area. Aluminum and iron do not appear as prone to heavy precipitation of insoluble phosphate minerals as calcium. The use of the two ions to improve bio-char phosphates adsorption is of considerable interest. This is especially important as both ions tend to create mildly acidic groups on the surface rather than the basic groups produced by ions such as calcium and sodium (Zhang et al., 2007).

It has been reported that acidic functional groups on the surface of bio-char can significantly increase nitrogen adsorption capacity. These acidic functional groups can be generated with oxidizing agents such as H2O (steam) and CO2 at high temperatures (300-700oC) or with ozone at room temperature (Chaing et al., 2002; Valdes et al., 2002). The development of low cost adsorbents such as: calcium carbonate, calcium, kaolinite, red mud, activated alumina, activated carbon from tamarind nut shell and bark, fly ash, and blast furnace slag, to remove phosphorous from aqueous streams has received increased attention in the last 10 years. In basic solutions (pH>7), biomass ashes with high contents of calcite are known to be efficient agents for the removal of phosphate. In acid conditions, aluminum and iron induce the precipitation of phosphates.

Biochar as a Soil Amendment

The use of biochar as a soil amendment has been investigated since the early 1800’s (Young, 1804; Lehmann et al., 2006; Lehmann and Joseph, 2009). Beneficial characteristics of biochar as a soil amendment are its high cation exchange capacity (CEC; 40 to 80 meq per 100 g), high surface area (51 to 900 m2 g-1), which leads to increased soil pH and water holding capacity, and affinity for micro- and macro- plant nutrients (Lehmann, 2007; Laird, 2008; Gaunt and Lehmann, 2008; Cheng et al., 2008; Novak et al., 2009; Lehmann and Joseph, 2009; Roberts et al., 2010).

Much of the interest in biochar as a soil amendment was prompted by studies of Amazonian soils (Terra Preta) where the presence of charcoal was associated with significant improvements in soil quality (e.g., soil organic matter [SOM] and nutrient concentrations) and increases in crop yields (Glaser et al., 2002). Early residents (450 BCE to 650 CE) of the Amazon River Basin applied large amounts of char and ash generated from cooking fires, plus other debris that included bones, pottery, and manure to soil (Sombroek, 1966). Improvements in chemical and physical properties of these soils have persisted for hundreds, if not thousands of years (Lehmann and Joseph, 2009; Lehmann, 2007; Novotny et al., 2009). Application of charcoal reportedly increases soil pH, and in acid soils charcoal decreases the concentration of Al, which often limits crop growth in the tropics (Hecht et al., 1982; Piccolo et al., 1997). When soil was amended with charcoal, annual crop yields increased by 100% or more (Steiner et al., 2007). In many tropical and subtropical soils, charcoal increases exchangeable bases, CEC, and nutrient availability, decreases soil density, and improves water holding capacity (Piccolo et al., 1997; Glaser et al., 2002; Liang et al., 2006; Busscher et al., 2010; Novak et al., 2010). The addition of biochar to soil also increases soil C concentrations that further improve nutrient storage and soil physical properties (Cheng et al., 2008).

Biochar has a market values due to its ability to absorb N and P from soils and waste water streams  and its potential to increase cation/anion exchange capacity. Any value added chemical processing treatment performed on the char to improve its nutrient adsorption capacity is will increase production costs. However, the goal of a further processing treatment is to increase the economic value by creating more uses or products providing a greater marginof return on the cost of production.

The study of biochar as a soil amendment in temperate zones of North America focuses on whether the benefits observed in the tropics can in the short term (i.e., one or two growing seasons) be reproduced in temperate climates. Observations to date have been inconsistent. In a sub-tropical Norfolk loamy sand (Fine-loamy, kaolinitic, thermic, Typic Kandiudult) in the southern USA, amendment with biochar at a rate of 2% increased pH, CEC, Ca, Cu, K, and P, but decreased N, Mg, Mn, Na, S, and Zn after a 67‑d incubation (Novak et al., 2009). In the same soil, C mineralization rates were negligible and did not increase after amendment and laboratory incubation; N immobilization occurred in as few as 25 d, with some N remobilized after 67 d (Novak et al., 2010). Char applications decreased N and P leaching but increased the leaching of K after a 500‑d laboratory incubation (Laird et al., 2009). Because bio-chars are significantly more stable than the fast and slow-cycling fractions of soil organic matter, the effects of bio-char additions to soil can have significant long-term benefits on soil C sequestration (Lehmann et al., 2006; Lehmann, 2007). The mean residence time (MRT), or turnover rate of C in charcoal-amended soil could exceed 1000 years (Swift, 2001). Biochar blended with switchgrass (Panicum virgatum) residue produced mixed outcomes on soil aggregation, rates of water infiltration, and water holding capacity (Busscher et al., 2010). Biochar-amended soils appear to emit less greenhouse gas than do soils amended with crop residues (Gaunt, 2008; Roberts et al., 2010).  Beneficial characteristics of biochar have been linked to the highly organized C-ring structure of biochar (Mykola et al., 2008; Keiluweit et al, 2010).  Biochar production protocols have yet to be standardized; therefore, any beneficial or detrimental effects on soil may vary with pyrolysis temperature and feedstock (Novak et al., 2010).

Joseph et al. (2009) has proposed a classification scheme based on these characteristics to describe the differences in the quality of biochar. We refer the reader to that review for specific details. A standard group of characterization tests has been suggested as biochar quality template (Lehmann and Joseph, 2009; McLaughlin et. al., 2009; Shinogi and Kanri, 2003).  Shinogi and Kanri (2003) have suggested physical characteristics (yield, surface area, bulk density) and chemical properties (total carbon, total nitrogen, pH, fixed carbon, ash content, and volatility) as indicators of biochar quality.  As data bases become available with these and other biochar characteristics (cation exchange capacity, absorbance) we will be able to address the issue of biochar quality and what characteristics are better suited for the desired goals of biochar applications  (Sohi et al., 2010).  To determine energy budgets for carbon mitigation policy or to assess the economic viability of amending soils with biochar, there must be evaluations of biochars produced from a variety of feedstocks, and analysis of the physio-chemical effects of biochar amendment at various rates, across climates, and among soil types. In Washington State (USA) about 15.4 million Mg per year of potential biofuel feedstocks are unused; these include forestry and agricultural residues, and animal manures (Washington State Dept. of Ecology, 2010).

The production, transport and application of biochar has some safety concerns of which users should be aware, however if precautions are taken they are manageable.  The primary concern for human and environmental risk is the particulate matter (PM).  Biochar dust and vapor fall between the U.S.A Environmental Protection Agency (EPA) air quality standards of PM2.5 and PM10 (Blackwell et. al., 2009; EPA, 2006).  These PM standards are associated with fine and ultra fine particles, which can contain volatile compounds both organic and inorganic.  During biomass pyrolysis several studies have shown the presence of  acetic acid, furylaldehyde, methyl acetate and several other volatiles (Amonette and Joseph, 2009; Greenberg et al., 2005).  Emissions of polycyclic aromatic hydrocarbons (PAH) have also been a concern however results from the literature have been mixed as pyrolysis temperatures below 700oC are not a major contributor to PAHs (www.biocharfarm.org; Verheijen et al., 2010; Garcia-Perez, 2008).  When dealing with these issues of volatiles and PM sizes individual biochars vary due to; temperature, pyrolysis type and feedstock influence each biochar its characteristics.  For example, rice husk biochar can contain crystalline material that could be toxic (Blackwell, 2009).  The exposure to the dust particles through inhalation appears to be the primary concern.  Verheijen et al. (2010) report that from the research the greatest concern for environmental pollutants and health concerns associated with biochar will come from small-scale pyrolysis units. In addition to volatiles the dust can remain in the air and be transported through air currents causing air quality issues in surrounding areas. As the EPA reviews its air agriculture quality standards in 2011 and as the use of biochar becomes more widespread the dust portion of the equation will need to be further addressed.

Outside of the health risks associated with the fine particles of biochar there is the added risk of fire hazard.  The United Nations recognized the hazard of spontaneous combustion for biochar giving it a Class 4.2 rating for transport.  The classification arises because of the nature of small particles in closed spaces to ignite similar to dust in grain elevators and /or coal dust in mines (Giby et al., 2007).  It is suspected that perhaps free-radicals are produced during pyrolysis attaching to the biochar surface may play a role leading to freshly produce biochar combustion (Verheijen et al., 2010; Amonette and Joseph, 2009).  In an attempt to counter act this hazard Blackwell et al. (2009) wetting, covering, and adding retardant chemicals to the biochar. However, the industry needs to develop standards to allow for safe cost efficient transport before expanding biochar use.

The transportation and application of biochar is still in the experimental process.  In the above section we addressed some of the risk of biochar dust and it ability for air transport.  Within our research at Prosser, WA pelletizing the feedstock prior to pyrolysis reduces dust production for easier transportation.  Simply putting biochar in large burlap bags during transportation has reduced biochar loss during transportation and application.

Methods of application for biochar in the field are currently being invested and are dependent on the purpose, crop and form of biochar.  Basic agronomy has addressed the application of fertilizers and provides a working base for the application of biochars.  Incorporation, banding, side-dressing, broadcast, foliar dusting, and liquid incorporation all possibilities; however one must take into consideration the precautions already discussed about health, safety and loss.  Blackwell et al. (2009) provides an excellent summary of these methods and there pros and cons.  We have tried several methods of hauling and application a summary is presented here. In our research biochar incorporation by rotovation was an acceptable method once the biochar was on the ground.  We found it difficult to apply the material on the field without loss to adjacent fields or plots unless there was no measureable wind.  Incorporation with a liquid (manure effluent or water) was difficult because in the early stages of mixing the hydrophobic nature of biochar did not allow good mixing. We had success with hand broadcasting to field plots with light hand rake incorporation immediately followed by 14 mm of irrigation.  In order for commercial growers to use this product a uniform method of effective application must developed with already commercial application equipment in mind.  Until this is addressed application of biochar will remain at the small scale level.

     As similarly discussed in previous sections temperature, feedstock, and pyrolysis technology play an important role in the individual characteristics of biochar.  Economic feasibility must be included in the cost of transportation, feedstock availability, energy (both used and supplied), production and application balanced with profitability.  The profitability factor also includes economic and environmental market tradeoffs.  The economics of biochar ultimately depends on the predominant use of the technology.  When or if the predominant end use of the pyrolysis technology changes, the economics of the whole system change also.  If higher C concentrations are desired then bio-oil production will most likely decrease.  Whether the use is in soil, for energy, or nutrient recovery individual there is a need for individual characteristics that favor the beneficial outcome.

A Novel Biochar Application

The integration of anaerobic digestion, pyrolysis, and P recovery in the dairy environment describes a natural place for the present research.  The existing energy infrastructure surrounding an anaerobic digester would allow for extra energy production through the capture of syngas.  This same infrastructure has the capability of drying down the anaerobically digested fiber through excess heat produced during digesting and power co-generation.  This drying down process is necessary if the material is to be pelletized.  Engineering of digesters brings the effluent flow to a single collection point source prior to fiber separation allowing for the production of biochar to be strategically aligned with P recovery.  The research presented in the following chapters is a proof of concept of using biochar for P recovery in a dairy setting that will ultimately spur on future engineering and precision. The following chapters represent a novel approach to addressing the use of anaerobically digested fiber to produce a biochar that is used as a low cost simple technological approach for P recovery from anaerobically digested dairy effluent.

The second chapter of this dissertation evaluates the production of biochar from 5 types of feedstocks and their effect on selected soil properties. A series of experiments was established to characterize biochars made from four feedstocks regionally available in the Pacific Northwest (wood pellets [Pseudotsuga menziesii (Mirb.) Franco], softwood bark [P. menziesii], switchgrass [Panicum virgatum] straw, anaerobically digested fiber) to determine their effect on select properties of five soils.  Soils were amended with 9.8, 19.5, and 39.0 Mg ha-1 of each of the four biochars and evaluated for changes in pH, water holding capacity, N mineralization, and change in soil C.  In response to the gap in scientific knowledge related to temperate soils our objectives were to: (1) characterize biochars made from two woody and two herbaceous feedstocks; and (2) determine their effects on soil at three rates of amendment.  Our research demonstrates that in temperate soils biochar feedstock may not be as important a variable as soil type for increasing C content and pH, but can influence N mineralization and WHC.

In the third chapter we evaluated the use of biochar produced from anaerobic digested fiber (APF) for capturing P from anaerobically digested dairy effluent.  To produce biochar a barrel retort slow pyrolysis unit was constructed.  P recovery potential of biochar was measured using a KH2PO4 P standard and anaerobic digested effluent (ADE).  Finally, a small scale filter system (1:25 biochar to ADE) was constructed to measure P reduction in the ADE.  Olsen P concentrations were analyzed to account for the plant available labile fraction of P on the biochar.  Liquid and solid state 31P nuclear magnetic resonance spectroscopy (NMR) was also utilized to characterize the P species on the biochar before and after use as a filtering media.   The objectives of this paper were to 1) characterize biochar created from ADF in a barrel retort, 2) characterize the P on the biochar before and after use as a filter media in ADE, and 3) quantify P.  Our central hypothesis is that the biochar will (1) sorbs phosphate from anaerobically digested dairy liquid manure effluent (2) reduce total P in lagoons and (3) that the sorbed P on the biochar complex will be in plant available forms.

Lastly, chapter four will determine the availability of biochar P and its effect the soil.  A series of soil incubations were set up to evaluate mineralization with three different rates of biochar application.  The objectives of this study were to 1) determine if the biochar and P affected the soil and, 2) determine if the coated biochar has significantly different effects on those properties.  The central hypothesis for this section the P biochar from the previous chapters will 1) increase C-mineralization rates, 2) decrease N-mineralization and, 3) increase P WXP and Olsen P levels in the soil.

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Figure 1: Phosphorus cycle taken from Pierzynski et al. 2005.

CHAPTER TWO

INFLUENCE OF BIOCHAR ON SOIL pH, WATER HOLDING CAPACITY, NITROGEN AND CARBON DYNAMICS

Jason Streubel, Harold Collins, Manuel Garcia-Perez, Julie Tarara, David Granatstein, Chad Kruger

(Accepted Soil Science Society of American Journal January 30, 2011)

Abstract

A series of experiments was established to characterize biochars made from four feedstocks regionally available in the Pacific Northwest (wood pellets [Pseudotsuga menziesii (Mirb.) Franco], softwood bark [P. menziesii], switchgrass [Panicum virgatum] straw, anaerobically digested fiber) to determine their effect on five soils.  Soils were amended with 9.8, 19.5, and 39.0 Mg ha-1 of each of the four biochars and evaluated for changes in pH, water holding capacity, N mineralization, and change in soil C.  The C content of biochars derived from the herbaceous feedstocks was 60 to 67 g g-1 whereas that of the woody feedstocks was >75 g g-1. In amended soils we found that biochars, regardless of origin, significantly raised the pH of all soil types 0.1 to 0.9 units with the greatest impact on a Quincy sand soil. The biochars increased soil C and water holding capacity at the higher rates of amendment depending on soil and biochar type. Nitrogen mineralization rates decreased in three of the five soils across all feedstocks.  There were significant correlations (r ³0.9) between biochar C measured and biochar C added regardless of feedstock or soil type.  Our research demonstrates that in temperate soils biochar feedstock may not be as important a variable as soil type for increasing C content and pH, but can influence N mineralization and WHC.

  Introduction

As world population increases so does the demand for global energy.  The grand challenge of meeting these requirements from fossil fuels alone is becoming increasingly difficult.  Utilizing alternative energy production methods is paramount to addressing the problem of energy security. The thermo-chemical conversion (pyrolysis) of plant biomass is being studied globally for its potential to produce second-generation liquid biofuels and the co-product charcoal, which can be used as a soil amendment. Biochar is currently the accepted term for pyrolysis-derived charcoal when designated for use as a soil amendment (Sohi et al., 2010); however char, charcoal, and black carbon also has been used.  Pyrolysis of biomass yields 60-70% of the original biomass as a bio-oil and 15-25% as biochar with the remaining 5-15% as ash (Graham et al., 1984; Czernik et al., 1994; Bridgwater et al., 1999; Mohan et al., 2006; Huber and Corma, 2007). Biochar is the carbon-rich product remaining after biomass has been heated at low temperatures (ca. 350°C to 600°C) in an environment with little or no oxygen (Amonette and Joseph, 2009).  The energy potentials from pyrolysis give cause for furthering research on biochar.

The use of biochar as a soil amendment has been investigated since the early 1800’s and has been recently rekindled due to concern about global energy shortages and climate change (Lehmann et al., 2006; Lehmann and Joseph, 2009). Beneficial characteristics of biochar as a soil amendment are its high cation exchange capacity (CEC; 40 to 80 meq 100 g-1), high surface area (51 to 900 m2 g­1), which leads to increased soil pH and water holding capacity, and affinity for micro- and macro- plant nutrients (Lehmann, 2007; Laird, 2008; Gaunt and Lehmann, 2008; Cheng et al., 2008; Novak et al., 2009; Lehmann and Joseph, 2009; Roberts et al., 2010). Much of the recent interest in biochar as a soil amendment was prompted by studies of Amazonian soils (Terra Preta) where the presence of charcoal was associated with significant improvements in soil quality (e.g., soil organic matter [SOM] and nutrient concentrations) and increases in crop yields (Glaser et al., 2002; Sombroek, 1966; Lehmann and Joseph, 2009; Lehmann, 2007; Novotny et al., 2009). Application of charcoal reportedly increased soil pH, and in acid soils charcoal decreases the concentration of Al, which often limits crop growth in the tropics (Hecht et al., 1982; Piccolo et al., 1997). When soil was amended with charcoal, annual crop yields increased by 100% or more (Steiner et al., 2007). In many tropical and subtropical soils, charcoal increases exchangeable bases, CEC, and nutrient availability, decreases soil density, and improves water holding capacity (Piccolo et al., 1997; Glaser et al., 2002; Liang et al., 2006; Busscher et al., 2010; Novak et al., 2010). The addition of biochar to soil also increases soil C concentrations that further improve nutrient storage and soil physical properties (Cheng et al., 2008).

Current research on biochar as a soil amendment focuses on whether the benefits observed in the tropics can in the short term (i.e., one or two growing seasons) be reproduced in temperate soils. Observations to date have been inconsistent. In a Norfolk loamy sand (Fine-loamy, kaolinitic, thermic, Typic Kandiudult) in the southern USA, amendment with biochar at a rate of 22.4 Mg ha-1 increased pH, CEC, Ca, Cu, K, and P, but decreased N, Mg, Mn, Na, S, and Zn after a 67‑d incubation (Novak et al., 2009). In the same soil, C mineralization rates were negligible and did not increase after amendment and laboratory incubation; N immobilization occurred in as few as 25 d, with some N remobilized after 67 d (Novak et al., 2010). Biochar applications decreased N and P leaching but increased the leaching of K after a 500‑d laboratory incubation (Laird et al., 2009).  Biochar incorporated with harvested switchgrass (Panicum virgatum) residue produced mixed outcomes on soil aggregation, rates of water infiltration, and water holding capacity (Busscher et al., 2010). The mean residence time (MRT), or turnover rate of C in charcoal-amended soil could exceed 1000 years giving the possibility of increasing soil C storage (Swift, 2001).

As biochar production protocols have yet to be standardized, any beneficial or detrimental effects on soil may vary with pyrolysis temperature and feedstock (Novak et al., 2010). To determine energy budgets for C mitigation policy or to assess the economic viability of amending soils with biochar, there must be evaluations of biochars produced from a variety of feedstocks, and analysis of the physio-chemical effects of biochar amendment at various rates, across climates, and among soil types including the temperate soils.  In Washington State (USA), 15 M Mg y-1 of potential biofuel feedstocks are currently unused; these include forestry and agricultural residues, and animal manures (Frear et al., 2005).  In response to the gap in scientific knowledge related to temperate soils our objectives were to: (1) characterize biochars made from two woody and two herbaceous feedstocks; and (2) determine their effects on soil at three rates of amendment.

Materials and Methods

Feedstocks and Production of Biochar

Four feedstocks were collected from sources in Washington state, USA: 1) Douglas fir wood pellets (WP; Pseudotsuga menziesii [Mirb.] Franco) distributed locally (Prosser, WA), 2) Douglas fir (softwood) bark (SB; P. menziesii; Swanson Bark & Wood Products, Longview, WA), 3) switchgrass straw (SW; USDA-ARS, Paterson, WA), and 4) animal-digested fibers (ADF) produced after processing dairy manure through a GHD  Plugged Flow (GHD Inc., Wisconsin) anaerobic digester (Outlook, WA). The biochar studied was produced in a batch pyrolysis reactor built by the Biological Systems Engineering Department at Washington State University (Granatstein et al., 2009) containing 200 g of biomass. The batch reactor vessel was a 585-mm long and 100-mm diam. horizontal metal tube (volume: 4.59 L) heated by a Lindberg/Blue M (Model HTF55322A) furnace. The reactor was heated to 500oC (heating rate: 25oC min-1) and kept at that temperature for 30 minutes. After each run the oven was turned off and the biochar was left inside the reactor to cool to ambient temperature. The pyrolysis vapors were evacuated from the reactor using 1 L min-1 (Normal Temperature and Pressure: NTP) of N2 as carrier gas.

Bulk soil samples were collected from the top 30 cm of five soil types representative of the cropping and agroclimatic regions of Washington State. Quincy sand (sand, mixed, mesic Xeric Torripsamment) is a young alluvial soil, with low nutrient and water holding capacity, occurring in central Washington and cropped in vegetable-based rotations under irrigation. Naff silt loam (fine-silty, mixed, mesic Ultic Argixeroll), Palouse silt loam (fine-silty, mixed, mesic Pachic Ultic Haploxeroll), and Thatuna silt loam (fine-silty, mixed, mesic Boralfic Argixeroll) are soils formed from loess deposits in the Palouse region of eastern Washington. These soils formed under grassland and are cropped to small grains under rain-fed conditions. Hale silt loam (coarse-loamy over sandy or sandy-skeletal, mixed, mesic Aquic Haplorthod) formed from loess and volcanic ash over glacial outwash in western Washington. Typical land uses are rain-fed row cropping including specialty horticulture crops, with some summer irrigation. Selected characteristics of the five soil types are listed in order of increasing soil C in Table 1.

Biochar and Soil Analyses

Each soil was amended with three rates (9.8 Mg ha-1 [0.4%]; 19.5 Mg ha‑1 [0.75%]; and 39.0 Mg ha-1 [1.5%]) of each biochar, with a non-amended soil for control.  Total soil organic C and N were determined by dry combustion on an elemental analyzer (CNS Vario EL III, Elementar, Hanua, Germany). The pH of each biochar and amended soil was determined using the method of Robertson et al. (1999) with 10 g soil or biochar and 20 mL of deionized water on a pH meter (Model 445, Corning Inc., Corning, NY). The CEC of amended soils was measured using the ammonium replacement method (Horneck, 1989) with 10 g of air-dried sieved (<2.0 mm) soil. Water retention curves of amended soils for calculating water holding capacity (WHC) were determined by pressure plate analysis at 0 and 0.1 MPa (Klute, 1986). Macro- and micro-elemental composition (P, K, S, B, Ca, Cu, Fe, K, Mg, Mn, and Zn) of biochars was determined by inductively coupled plasma atomic emission spectroscopy (ICP; DV3300, PerkinElmer, Waltham, MA) according to Isaac and Johnson (1998).Soil N mineralization potentials were determined over a 49-d incubation. Amended soils (10 g) were incubated at 70% of field capacity at 25oC and extracted weekly with 50 mL of 2 M KCl, shaken for 1 h on a rotary shaker, then filtered through a Type A/E glass fiber filter (Gelman Sciences Inc., Ann Arbor, MI). The extracts were analyzed colorimetrically for NO3‑N and NH4-N using a flow-injection analyzer (FIA) equipped with an auto sampler (QuikChem AE; Lachat Zellweger, Loveland, CO).

Carbon mineralization was used as a measure of soil biological activity and the stability of biochar C using the static incubation method (Zibilske, 1994). Soil samples with biochar (25 g) at each biochar application rate were adjusted to 70% FC in 160-mL containers and were incubated at 25°C for 30 weeks. Headspace CO2 was measured weekly by direct injection of gas samples into an infrared gas analyzer (IRGA; type 225- MK3, Analytical Development Co., Hertfordshire, UK). After each analysis, samples were adjusted for moisture loss and the headspace was returned to ambient by degassing with compressed air.

Carbon sequestration potentials of biochar-amended soils were calculated from rates of C mineralization. The [CO2] evolved during C mineralization was used to determine the size of the functional pools of soil C for each amended soil (Collins et al., 1999; Paul et al., 2003). The size and turnover rates of each pool were determined by curve fitting the estimates of CO2 evolved per unit time (Ct), using a three-component first-order model:

Ct = Cae-kat + Cse-kst + Cre-krt

where Ct represents total C, the subscripts a, s, and r represent the active, slow, and resistant pools, respectively, and k is the decomposition rate constant for each pool. Three parameters, the size of the active pool (Ca), and the decomposition constants of the active (ka) and slow (ks) pools, were estimated using nonlinear regression (Systat, Inc., Evanston, IL). The slow pool is defined:

Cs = Ct – Ca – Cr.

The Cr, or resistant organic fraction, is assumed to contribute negligibly to CO2 efflux during laboratory incubation (Paul et al., 2006). The size of this pool was determined by acid hydrolysis using 1 g of amended soil in 6 M HCl for 18 h. Digested samples were washed three times with deionized water, oven dried at 55oC, and ground to pass through a 180-μm screen. The C content of the sample was determined by dry combustion as described above.  Mean residence time is the reciprocal of the decomposition rate constant (k-1) derived from the first order rate reaction at laboratory incubation temperature (25ºC).

Statistics

   Data were tested for normality by Shapiro-Wilk’s W test and for heterogeneity of variance by Levene’s test using SAS (version 9.2, SAS Institute, Cary, NC). Associations between feedstock and nutrient analysis, pH, WHC, and C and N mineralization were assessed using PROC MIXED and ANOVA with pairwise comparisons by Tukey at P < 0.05 (version 9.2, SAS Institute, Cary, NC).

Results and Discussion

Feedstock C, N, and S

The C concentrations of the pre-prolysis feedstocks ranged from 480 to 432 g kg-1 with ADF the highest and SW the lowest (Table 2). The N and S concentrations of each feedstock differed by plant type: both N and S concentrations were higher in the herbaceous (average 21.8 g kg-1 N; 2.3 g kg-1 S) than in the woody feedstocks (2.3 g kg-1 N; 0.3 g kg-1S). The values of C and N for ADF (Table 2) are similar to those reported for manure fiber (Cao and Harris, 2010). The C:N of the woody feedstocks (142 to 398) was higher than that of the herbaceous feedstocks (18 to 24), which can be accounted for by the difference in hemicellulose, cellulose, and lignin concentrations between woody perennials and grasses (Phyllis Database for Biomass and Waste, 2010; Cherney et al., 1988). The lignin content of various softwoods and barks varies between 20 and 50% of dry mass (Phyllis Database for Biomass and Waste, 2010) compared to 7 to 9% for switchgrass (Cherney et al., 1988); cellulose concentrations average 0.6% of dry matter in woody species and 36% in herbaceous species. The N concentration of switchgrass ranged from 10 to 30 g kg-1 whereas both tree bark and wood contained on average 3.1 g kg-1.

 

Elemental Concentrations of Biochars

The SW and ADF biochars had C concentrations of 0.592 and 0.658 g g-1, respectively, whereas SB was 0.727 g g-1 and WP was 0.782 g g-1 (Table 2). The biochars from herbaceous feedstocks had higher N concentrations than the woody biochars, similar to analyses of the feedstocks. Loss of C during pyrolysis is related to the original feedstock concentrations of ash, lignin, cellulose, and hemicellulose (Demirbas, 2006). The SW and ADF biochars lost 45 and 48% of their original feedstock C content, respectively, in contrast to the woody biochars (SB and WP) that lost 39 and 40% during pyrolysis. Nitrogen concentrations were 2% for the herbaceous biochars and <0.4% for the woody biochars. The SW and ADF biochars respectively lost 66 and 58% of their original N content following pyrolysis, compared to 59% for the SB biochar and 62% for the WP biochar (Table 2). Yang et al. (2006) and Güllü and Demirbas (2001) found that hemicellulose pyrolyzed between 220 and 315oC, and cellulose between 315 and 400oC; however, lignin was difficult to pyrolyze under low oxygen concentrations. Cao and Harris (2010) suggested that the decrease in N can be attributed to volatilization during heating and that some of the N-containing structures in the biochar (e.g., amino acids, amines, amino sugars) were condensed into recalcitrant forms (Shinogi and Kanri, 2003).

Acid hydrolysis of the biochar removed labile forms of C such as simple sugars, fats, and oils that may remain after pyrolysis. Herbaceous biochar materials (SW, ADF) lost about 8% of their total C after acid hydrolysis, whereas biochars from the woody feedstocks (SB, WP) remained largely unchanged (Table 3). The additional C loss may have originated from condensates of the bio-oil that coats the biochar following pyrolysis. Under low temperatures (250 to 600oC), the thermo-chemical reaction between pyrolysis and liquefaction can be variable, leaving tar or bio-oil on the surface of the biochar from the unstable intermediate fragments (free phenoxyl radicals) remaining between pores in the solid structure (Demirbas, 2000).

Ash contents were higher in the herbaceous (15 to 21%) than woody biochars (1 to 5%), following the order of SW > ADF > SB > WP. Shinogi and Kanri (2003) reported the same trend of higher ash contents in herbaceous feedstocks, concluding that C, N, H, O, and S are lost during heating as the inorganic salts that comprise the ash are not volatilized. Macronutrient concentrations varied by plant type and were higher for the herbaceous biochars (Table 4). Phosphorus concentrations were 10 times higher in the SW and ADF than in SB and WP biochars. The ADF tended to have the highest concentrations of micronutrients, except for Mn which was higher in SB biochar. Structural components of the softwood bark feedstock may explain the higher concentration of Mn: it is an essential nutrient for the construction of lignin and for the synthesis of phenolic acids, both of which are in high concentrations in tree barks (Havlin et al. 2005). Altland et al. (2008) reported concentrations of Mn that ranged from 5 to 30 mg L-1 in softwood bark.Among herbaceous feedstocks, concentrations of all elements measured except K were higher in the ADF than in the SW biochar. The ADF feedstock was obtained from a dairy with a flush system for managing manures from housing and milking barns, and produced by an anaerobic digester to produce methane gas, which likely leached soluble elements such as K from the ADF feedstock (Hart et al., 1997; Downing et al., 2010). The concentrations of B, Cu, and Zn were 2, 8, and 20 times higher in ADF, respectively, than in the SW biochar, most likely resulting from mineral supplements in the dairy cattle feeds and the Cu foot baths used to treat hoof infections. Cao and Harris (2010) found similar elemental concentrations and trends in dairy manure; however, they did not measure Cu concentration.

Influence of Biochar Additions on Soil Properties

Soil pH

The alkaline nature of the biochar exchanges H+ with the surrounding soil, causing a rise in soil pH. Soil pH increased 0.4 to 0.8 units for soils amended with 39 Mg ha‑1 of the herbaceous biochars and by 0.1 to 0.4 units for amendment with woody biochars at the same rate (Figure 1).  The Quincy sand exhibited the greatest and most rapid increase in pH of all soil types. This response differed from the silt loam soils which showed smaller increases in pH from 9.8 to 19.5 Mg biochar ha-1,with a further increase at 39 Mg biochar ha-1.

The pH of tropical and temperate soils has been shown to vary in response to biochar additions.  The amount of increase is dependent on the biochar types and the buffering capacity of the soil, meaning its ability to resistance changes in pH which depends on several factors including the soil’s organic and mineral content, and the soil’s physical properties (e.g., density and particle size; Havlin et al., 2005).  The pH of the herbaceous biochars was two units higher (9.4) than woody (7.4) biochars, due to higher ash concentrations in these feedstocks (Table 4). At 300 to 600oC, chemical cracking of hemicellulose and cellulose occurs, as pyrolysis temperatures rise, alkali salts separate and raise the pH of the material (Abe et al., 1998; Shinogi and Kanri, 2003). The inherently lower buffering capacity of sands compared to silt loam and clay soils may explain the rapid increase in pH.  Cox et al. (2001) reported for a Palouse silt loam soil in Washington State an increase in soil pH of 0.8 units after the application of 110 Mg ha-1 of fly ash obtained from a local gasification power plant operating between 800 and 1000oC. In a tropical soil, Steiner et al. (2007) found amendment with 11 Mg ha-1 of wood charcoal was not sufficient to significantly alter soil pH. In an Atlantic coastal plain soil (USA), a 22.4 Mg ha-1 application of biochar increased soil pH from 5.9 to 7.7 (Novak et al., 2009). Yamato et al. (2006) raised the pH of three soils in south Sumatra by 1.0 to 1.5 units with an application of 37 Mg ha‑1 wood bark biochar. In a Columbian soil amended with 11 Mg ha-1 of a wood-derived biochar from a high temperature gasification process, pH increased by 0.5 units (Rodriguez et al., 2009). Rondon et al. (2006) also reported a pH increased of 0.4 units after addition of 100 Mg ha-1 (9%) of biochar amendment.

Water Holding Capacity

The effect of biochar on WHC was variable among soil types and biochars (Table 5). In the silt loam soils, significant increases in WHC at 0.1 MPa (1 bar) were consistently observed for the SW biochar, averaging 13.7% at the highest rate of amendment.  The SB biochar also increased WHC in the silt loam soils, but the difference was significant only at higher rates of amendment in the Thatuna and Hale series. Soil WHC increased significantly with WP amendment in Thatuna (by 12.5% at 39 Mg ha-1) and Hale series (by 34.2% at 39 Mg ha-1), but showed no effect in the Naff or Palouse silt loams.  The differences between the silt loams are most likely related to clay content.  The Hale and Thatuna soils have high clay contents as indicated by the presence of a Bt horizon or redox (NRCS Soil Survey, 2010).  Clay-enriched soils naturally have a higher water holding capacity and with biochar surface areas similar to those of clays (51 to 900 m2 g-­1), the addition of biochars to soil could cause the increase in WHC (Downie et al., 2009).  We suggest that the physical structure of the biochar could account for differences in the higher clay content soils (Downie et al., 2009; Canessa et al., 1994).  Silt loam soils typically hold 0.15 to 0.22 cm water cm-1 soil. The observed increases in WHC are equivalent to an additional 0.02 to 0.03 cm water cm-1 soil. There was no response to biochar amendment in the Quincy sand due to high variability in soil WHC, attributable to the single-grained structure of this soil and variable distribution of the biochar particles.

The reported effect of biochar on WHC has been mixed and is related to both biochar properties and soil type. The change in WHC in a Norfolk loamy sand(Fine-loamy, kaolinitic, thermic, Typic Kandiudult)soil also varied by biochar feedstock after a 28-d incubation (Novak et al., 2009). They found that a poultry-litter biochar decreased WHC while peanut hulls and pecan shells raised it only slightly; however, switchgrass biochar raised WHC by 16%. By contrast, also in a Norkfolk loam soil, Busscher et al. (2010) found no significant change in WHC at 11 Mg ha-1 amendment, but speculated that higher rates of amendment may increase WHC. Our results show a similar variance in results among amendment rates and soil types, but support the findings of Busscher et al. (2010).

Soil N Mineralization

Nitrogen mineralization varied among biochars and soil types. There was a consistent and significant decrease in N mineralized with the addition of biochar regardless of feedstock in the Palouse, Neff, and Thatuna soils. The N mineralized in the Hale silt loam increased significantly at the highest amendment rate of the SW biochar and a decrease for SB.  Quincy sand showed a significant increase in N-mineralization with application of 9.8 Mg ha-1 of ADF and with 39.0 Mg ha-1 of WP (Figure 2). Lower N mineralization may be related to the biochars’ concentration of volatile material that inhibits microorganisms in the soil and thus decreases NO3 production (Deenik et al., 2010).

The N mineralized early in the incubation was higher in all soils amended with the SW, ADF, and SB biochars than in the un-amended control (data not shown). Switchgrass residue added with biochar to a sandy soil could potentially stimulate the availability of mineralized N (Novak et al., 2009). Biochar amendment decreased N availability in some Amazonian soils (Lehmann et al., 2003), which may be beneficial since N fixation by biochar would reduce losses by leaching. Others found as much as 70% less N leached from soils that had been amended with biochar (Laird, 2008; Cheng et al., 2008). The addition of biochar (33.6, 67.2, and 100.8 Mg ha‑1) to soil increased biological N fixation in beans inoculated with Rhizobium spp., but N uptake was decreased, suggesting that immobilization occurred (Rondon et al., 2006).

Soil C

The Quincy sand soil showed the greatest increase in soil C due to its low initial native C content compared to the silt loam soils (Table 1).  The linear relationships between total C and C added as biochar indicate that for the soils tested; nearly all of the added biochar-C was accounted for in the total C analysis (Figure 3). We recorded a consistent increase in soil C but not a consistent doubling for all soils except for the Quincy sand (Supplemental Tables 1-5). We recovered less biochar C from the Quincy sand (Figure 3) than from the silt loam soils, also associated with the single-grained structure of the sand soil and the variable distribution of biochar particles in the soil (Kurth et al., 2006). At the higher rate (39.0 Mg ha-1) of biochar addition, soil C in the Quincy sand averaged a 150% increase, whereas the silt loams had average C increases of 25, 38, 55, and 63% for Hale, Thatuna, Palouse, and Naff soils, respectively (Supplemental Tables 1-5). Among the silt loam soils, as native soil C increased, the effect of biochar-C addition was diluted by the increasing background concentration. With increasing amendment rate, biochar derived from the woody feedstocks increased soil C in the Quincy sand more so than did the herbaceous biochars, with a two-fold increase in total C at 39 Mg ha-1 WP biochar (Supplemental Table 1). These increases are attributed to the higher C content of the woody than herbaceous biochars (Table 2).

We found a similar response for increases in the stable soil C pool (Cr) in biochar-amended soils (Figure 3). The Cr comprised between 50 and 90% of total soil C depending on soil type (Supplemental Tables 1-5). The Cr has been used to represent the persistence of C in soil (Paul et al., 2001, 2006).  Among all soil types and biochars, estimates of Cr were on average 80% of the total soil C pool after amendment (R2=0.993, ±SE 1.23). Kurth et al. (2006) also showed a strong correlation (R2 = 0.99) between the amount of C added as charcoal and the amount estimated by a peroxide/nitric acid digestion method.

C Mineralization

The C-mineralization coefficient, the percentage of total organic C evolved as CO2, has been used to determine the effect of soil and cropping systems management on the decomposition and cycling of SOM (Collins et al., 1992, 1999; Paul et al., 1999, 2006). Long-term incubations (>200 d) have also been used to differentiate functional C pools in soil (Motavalli et al., 1994; Paul et al., 1999; Cochran et al., 2007; Collins et al., 2010).

The rates of C mineralization that we observed among biochars in 225-d incubations are similar to those reported in other studies (Bruun et al., 2008; Zimmerman, 2010; Novak et al., 2010). The C‑mineralization rate curves were similar for the five amended soils (see example for SW biochar in Figure 4).  Carbon mineralization in each soil type was significantly greater in the initial days of the incubation which can be partly explained by the presence of labile C on the biochars (Smith et al., 2010). The increase in CO2 evolved was 1 to 10 mg CO2-C kg-1 soil d-1, depending on the feedstock. As mentioned above, under low pyrolysis temperatures (250 to 600oC), the thermo-chemical reaction between pyrolysis and liquefaction can be variable (Demirbas, 2000).  At 500oC, Keiluweit et al. (2010) found that the volatile matter content of biochar was still fairly high in fescue but had dropped significantly for Pinus ponderosa. Smith et al. (2010) showed that newly produced biochar added to soils increased C-mineralization for the first six days of incubation, and then rates returned to pre-addition levels. Using δ13C analysis they also showed that the increase in CO2 efflux originated from biochar-C and not from the native SOM.

Carbon mineralization after biochar additions followed the pattern of Thatuna>Palouse>

Hale>Naff> Quincy, a result of greater initial soil C in the silt loams than in the sand. The rate of biochar addition had only minor effects on total C mineralized (Supplemental Tables 1-5). Although the Hale soil had the highest soil C, it had the lowest proportion of C mineralized. This is likely a function of the higher clay content and Cr pool in this soil type, which inhibits SOM oxidation and microbial degradation. The percentage of total C mineralized decreased as the rate of biochar addition increased, similar to that observed by Bruun et al. (2008) and Zimmerman (2010). The reduction in C-mineralization results from dilution of soil organic C with C that is largely biologically inert.

The proportion of soil organic C in the active soil C pool (Ca) of the un-amended soils ranged from 0.2 to 1.4% with the silt loam soils containing higher concentrations of labile C. The Ca as a percentage of the total C declined nearly 40% among soil types as the rate of biochar amendment increased. Laboratory MRTs of Ca among soils ranged from 11 to 30 d in the un-amended soils and tended to decrease for all soils as application rates increased, except for the Quincy sand and Naff silt loam which showed >50% increase in MRT for all biochars as application rate increased (data not shown).

The proportion of total C in the slow C pool (Cs) ranged from 26 to 50% in the un-amended soils. The Cs, as a percentage of total C, declined 78, 48, 46, 32, and 28% for the Quincy, Naff, Palouse, Thatuna and Hale soils, respectively, as the rate of biochar amendment increased to 39.0 Mg ha-1. These reductions resulted from dilution of soil organic C, with the un-reactive biochar C comprising the majority of Cr after amendment (Supplemental Tables 1-5). Novak et al. (2010) reported similar findings of C mineralization when biochar was added to soil during a 67‑d incubation. Deenik et al. (2010) found that biochar with higher volatile mater increased mineralization rates in the first days of incubation. They suggested that biochars pyrolized at higher temperatures (>350oC) contain lower volatile concentrations on the surface of the biochar and therefore do not stimulate microbial populations as readily as biochars pyrolized at lower temperatures. They showed higher rates of respiration in the soil when high‑volatile biochar was added to the soil (Novak et al., 2010; Deenik et al., 2010).

The proportion of total C contained in the resistant soil C pool (Cr) comprised 50% of the total soil organic C in the un-amended Quincy sand to 90% after addition of 39.0 Mg biochar ha‑1 (Supplemental Tables 1-5). There were only minor differences due to the type of biochar. For the Quincy sand the mass of C in Cr increased ~1.2, 2.5, and 4-fold after addition of 9.8, 19.5, and 39.0 Mg biochar ha-1, respectively. For the silt loam soils the change in mass of C in Cr decreased in the order 1.3, 0.75, 0.7, 0.5-fold for the Naff, Palouse, Thatuna, and Hale soils, respectively, with minor effects due to the origin of the biochar. The MRT of Cr is between 1000 and 2500 y for surface silt loam soils of the midwestern US (Paul et al., 2001). Following a 3.2‑yr incubation, Knoblauch et al. (2010) estimated a MRT of >2000 years in soils amended with black carbon, and found a significant decrease in C mineralization in four soils compared with the same soils amended with rice hulls. These long-term studies indicate the stability of biochar in contrast to short-term outcomes previously published (Novak et al., 2010).

Conclusion

Charcoal has been used as a soil amendment in humid equatorial climates for centuries. Biochar amendment has the potential to improve temperate soils by adding C, raising pH, and increasing WHC. Biochar is an alternative end-product for local farming and forestry residues such as dairy waste, forage residues, forestry slash, and other production debris. Amendment of agricultural soils with biochar pyrolyzed from herbaceous and woody feedstocks had variable effects on soil properties depending on soil type, biochar feedstock, and rate of amendment. In the Hale silt loam and Quincy sand, rates of N mineralization showed an increase among some rates and biochar sources, whereas in the remaining silt loams N mineralization tended to decrease. In Quincy sand, Palouse, Thatuna, Naff and Hale silt loams amended with biochar, the percentage of total C mineralized decreased as the rate of biochar amendment increased. There was a linear relationship between the rate of biochar amendment and the change in C content of the soil. Soil pH increased with biochar amendments among all soil types and biochar feedstocks. The use of biochar to raise pH may be beneficial where long-term fertilizer applications have acidified soils. Conversely, carbonate-rich soils with high pH that are common in arid regions may not benefit from additions of biochar because increases in soil pH can be detrimental to micro-nutrient availability and to crop production. Soil water holding capacity varied among rates of amendment and biochar feedstock. In irrigated systems, appropriate choice of biochar feedstock could increase soil WHC and thus reduce either the frequency or amount of irrigation. The recalcitrant nature of biochars can also improve C sequestration in agricultural soils because the proportion of total C in biochar that is recalcitrant leads to long MRT. We found that biochar feedstock is not a significant factor in raising pH or C in five regional soils. However, further research is needed to better understand the effect of feedstock on WHC and N mineralization.  Field-scale research is also needed to understand interactions between biochar and the soil matrix.

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Table 1. Selected characteristics of the five soil types used in the laboratory incubations.

Soil Soil Characteristics
Series Texture C N S C:N C:S pH CEC
 ————— g kg-1 —————— cmol kg-1
Quincy Sand   4.3 (0.5) 0.5 (0.1) 0.2 (0.03) 8.6 22 7.1   3.3
Naff Silt loam 18.0 (1.0) 1.5 (0.1) 0.2 (0.02) 12.0 90 4.5 15.4
Palouse Silt loam 23.2 (0.5) 2.0 (0.1) 0.4 (0.11) 11.6 58 4.6 16.0
Thatuna Silt loam 26.9 (0.5) 2.4 (0.1) 0.4 (0.11) 11.2 67 4.4 16.1
Hale Silt loam 39.9 (0.9) 3.4 (0.1) 0.6 (0.10) 11.7 67 4.7 16.6

Std. deviation in parentheses.

Table 2. Change in C and N concentrations of feedstocks following pyrolysis at 500oC.

Biochar Carbon
Biochar Source Original FeedstockC content Char Yield Char C Concentration Char C Content Loss of C from Original Feedstock
g kg-1 g g g-1 ——- g kg-1 ——- %
SW 432 (8) 400 0.592 236.8 195.2 45.2
DF 480 (2) 380 0.658 250.0 231.0 48.0
SB 470 (2) 390 0.727 283.5 186.5 39.7
WP 477 (2) 360 0.782 281.5 195.5 41.0
Biochar Nitrogen
Original Feedstock N content Char Yield Char N Concentration Char N Content Loss of N from Original Feedstock
g kg-1 g g g-1 ——- g kg-1 ——- %
SW 23.5 (0.3) 400 0.0199 8.0 15.5 66.1
ADF 20.0 (0.2) 380 0.0223 8.5 11.5 57.6
SB   3.3 (0.1) 390 0.0035 1.4 1.9 58.8
WP   1.2 (0.1) 360 0.0013 0.5 0.7 62.2

Char yield – mass remaining following pyrolysis of 1 kg of feedstock.Std. deviation in parentheses.

Table 3. Carbon and N concentrations, and changes in concentrations after acid hydrolysis, for four biochars pyrolized at 500oC.

Pre-hydrolysis Post-hydrolysis % change
Biochar Source C N C N C N
———————- g kg-1  ———————–
SW 592 (13) 19.9 (0.7)   641 (12) 19.9 (0.3) 8.3 0
ADF 658 (10) 22.3 (0.2)   709 (17) 23.4 (0.4) 7.8 4.9
SB 727 (17) 3.5 (0.2)   737 (17)    3.5 (0.3) 1.4 0
WP 782 (18) 1.3 (0.2)   785 (15)    1.4 (0.1) 0.4 7.6

6 N HCl, 18 h. Std. deviation in parentheses.

Table 4. Elemental composition and pH of four biochars and a commercial activated charcoal control used in laboratory incubations.

Elemental Composition
Biochar Source Ash P K S Ca Mg Fe B Cu Mn Zn pH
  ———————————– g kg-1 biochar ———————————— —————— mg kg-1 biochar —————-
SW 208.1 (8) 4.7 (0.14) 32.8 (2) 1.1 (0.02)  8.7 (0.41) 4.6 (0.29) 0.62 (0.1)   63.8 (2)    7.7 (1.3) 108.8 (3)   33.7 (2) 9.4
ADF 154.4 (1) 7.6 (0.32) 11.7 (1) 3.0 (0.02) 24.0 (0.5) 7.0 (0.22) 1.28 (0.02) 148.8 (4) 162.9 (5) 184.0 (4) 230.0 (5) 9.3
SB 54.0 (2)   0.47 (0.03) 1.0(0.1) 0.23(0.01) 10.7 (0.1) 0.48 (0.02) 0.70 (0.03)   56.7 (8)    6.8 (0.7) 266.7 (25)   40.9 (2) 7.6
WP 11.6 (1)   0.22 (0.03) 1.0(0.1) 0.17(0.02) 2.0 (0.03) 0.30 (0.04) 0.25 (0.1)   91.2 (5)    3.1 (0.2)   93.3 (13)   29.1 (3) 7.2
AC 20.3 (3)   0.25 (0.01) 9.3 (0.3) 0.17 (0.01) 0.6(0.04) 0.28 (0.01) 0.04 (0.0) 110.3 (9)    9.2 (0.4) 0     0.7 (1) 9.6 (0.1)

SW-switchgrass; ADF- animal digested fiber; SB- softwood bark; WP- wood pellets; AC- activated charcoal.

Std. deviation in parentheses.

Table 5.Water holding capacity (WHC) of five soils at 0.1 MPa after variable rate amendment with four biochars pyrolized at 500oC.

Soil Series
Source BiocharRate Quincy Sand NaffSilt loam PalouseSilt loam ThatunaSilt loam HaleSilt loam
  Mg ha-1 —————————– WHC (cm cm-1) —————————–
SW 0 4.0 (0.3) 17.2 (1.2) 19.5 (0.2) 18.4 (0.6) 23.8 (1.7)
  9.8 4.4 (0.6) 17.4 (1.0) 19.9 (0.3) 20.7 (0.1)* 25.0 (0.6)*
19.5 4.3 (0.3) 19.5 (0.2)* 23.2(0.7)* 20.9 (0.1)* 25.0 (0.1)*
39.0 4.5 (0.3) 19.8 (0.2)* 23.5(0.8)* 20.7 (0.1)* 25.4 (0.7)*
ADF 0 4.0 (0.3) 17.2 (1.2) 19.5 (0.2) 18.4 (0.6) 23.8 (1.7)
  9.8 4.4 (0.4) 17.2 (1.1) 19.7 (0.1) 18.5 (0.4) 25.3 (1.1)
19.5 4.5 (0.4) 17.7 (1.4) 19.7 (0.1) 18.3 (0.7) 25.4 (0.9)
39.0 4.4 (0.5) 17.9  (1.4) 19.9 (0.1) 18.6 (0.1) 25.3 (0.7)
SB 0 4.0 (0.3) 17.2 (1.2) 19.5 (0.2) 18.4 (0.6) 23.8 (1.7)
  9.8 4.2 (0.2) 19.6 (1.1)* 22.6(0.8)* 20.1 (1.1) 24.8 (0.4)
19.5 4.0 (0.4) 18.0 (0.2)* 22.4(1.0)* 20.3 (1.0)* 23.4 (0.6)
39.0 4.0 (0.3) 19.9 (0.4)* 22.4(0.9)* 20.7 (1.3)* 26.2 (0.8)*
WP 0 4.0 (0.3) 17.2 (1.2) 19.5 (0.2) 18.4 (0.6) 18.4 (0.6)
  9.8 4.5 (0.5) 17.3 (0.8) 19.5 (0.1) 20.8 (0.3)* 24.8 (0.7)*
19.5 4.2 (0.3) 17.9 (1.3) 19.5 (0.1) 20.5 (0.1)* 24.7 (1.1)*
39.0 4.0 (0.1) 17.7 (1.4) 19.7 (0.2) 20.7 (0.2)* 24.7 (0.4)*

Std. deviation in parentheses.  Asterisks indicate significant differences from the un-amended soil at P =0.05.

 

Figure 1. Influence of biochar amendment rate on soil pH for five soil types and four biochars. Error bars are standard error of the mean and values plotted at zero is the pH of the soil prior to incubation. All biochar rate increases caused a significant increase in soil pH at p< .05 except as noted.

Figure 2. Total N-mineralization at 49d of laboratory incubation as a function of soil type and biochar.  Letters represent significant differences at p .05 Tukey’s pairwise comparisons.

Figure 3. Relationships between the recovery of biochar C (a) and acid

hydrolyzed biochar C (b) by the biochar C initially applied to each soil.

  

Figure 4. Rates of C mineralization of soils amended with switchgrass biochar for Quincy sand, and Naff, Palouse, Thatuna, and Hale silt loams measured during a 225-d laboratory incubation. Biochar amendment rates were 0, 9.8, 19.5, and 39.0 Mg ha-1.

Supplement Table 1.  Total C, C-mineralization and estimated pool sizes of the active (Ca), slow (Cs) and resistant (Cr) C pools for Quincy sand soil amended with four rates of biochar pyrolized from each of four feedstocks.

C increase C                      Mineralized Active Pool Slow Pool                             Resistant Pool
Soil/Biochar CharRate Total C fromnative CumulativeCO2-C C-min /TC Ca Cs Cr Cr/TC
g kg-1 g kg-1 % mg kg-1 % mg kg-1 g kg-1 g kg-1 %
Quincy 0   4.5 (0.4) a 0   252 (15) a 5.6 19 a 2.4 2.1 (0.1) a 46.9
 SG 4.0   4.6 (0.3) a 1.6 258 (6) a 5.6 38 b 2.7 1.8 (0.2) a 40.4
7.5   8.0 (0.9) b 77.4  270 (13) a 3.4 50 c 2.9 5.0 (0.8) b 63.0
15.0 11.0 (0.9) c 143.4  337 (15) b 3.1 86 d 2.9 8.1 (1.0) c 74.0
ADF 4.0   4.7 (0.6) a 3.0    270 (10) a 5.8 40 b 2.0 2.6 (0.4) a 57.0
7.5   4.9 (0.3) a 7.5 286 (6) b 5.9 49 bc 1.2 3.6 (0.3) d 74.5
15.0   9.9 (0.7) b 118.4   361 (19) c 3.7 72 d 2.0 7.8 (1.5) c 79.3
SB 4.0   5.7 (0.7) a 25.9   251 (11) a 4.4 28 b 3.3 2.3 (0.3) a 40.3
7.5   9.1 (0.8) b 100.3 258 (6) a 2.9 33 b 3.7 5.3 (0.3) b 58.3
15.0 12.3 (0.4) c 172.6   292 (17) b 2.4 43 b 3.1 9.2 (0.8) c 74.9
WP 4.0   5.6 (0.7) a 24.4 242 (6) a 4.3 18 a 1.6 3.9 (0.5) d 70.3
7.5 10.6 (0.9) b 134.0  248 (11) a 2.3 23 a 3.6 6.9 (0.8) e 65.6
15.0 13.6 (1.4) c 200.9  245 (14) a 1.8 27 a 1.7 11.9 (2.0) c 87.3

†Values in parentheses are standard deviations values followed by the same letter are not significantly different at P=0.05.  Abbreviations: TC—Total Carbon content; SW-switchgrass; ADF- animal digested fiber; SB- softwood bark; WP- wood pellets.

Supplemental Table 2.  Total C, C-mineralization and pool sizes of the active (Ca), slow (Cs) and resistant (Cr) C pools for Naff silt loam soil amended with four rates of biochar pyrolized from each of four feedstocks.

C increase CMineralized Active Pool Slow Pool                             Resistant Pool
Soil/Biochar CharRate Total C fromnative CumulativeCO2-C C-min /TC Ca Cs Cr Cr/TC
g kg-1 g kg-1 % mg kg-1 % mg kg-1 g kg-1 g kg-1 %
Naff 0 18.1 (1.0) a 0 1645 (31) a 9.1 39 a 6.5 11.6 (0.6) a 64.1
 SG 4.0 19.9 (0.4) a 10.2  1690 (162) a 8.5 65 b 4.7 15.2 (1.0) b 76.4
7.5 22.6 (1.0) b 25.0 1548 (84) a 6.8 118 c 4.0 18.6 (1.6) c 82.1
15.0 27.7 (0.6) c 53.5 1407 (14) b 5.1 167 c 5.5 22.3 (1.9) c 80.4
ADF 4.0 20.7 (1.0) a 14.2  1658 (108) a 8.0 79 b 5.8 14.9 (1.4) b 72.1
7.5 22.9 (0.6) b 26.5 1908 (21) b 8.3 54 b 5.4 17.5 (0.6) c 76.4
15.0 26.6 (0.4) c 46.7 2047 (64) c 7.7 47 b 4.8 21.7 (3.3) c 81.8
SB 4.0 21.4 (0.3) a 18.0  1717 (63) ab 8.0 52 b 5.8 15.6 (0.9) b 72.8
7.5 23.0 (0.7) b 27.2 1800 (19) b 7.8 32 a 4.8 18.2 (0.7) c 79.2
15.0 30.2 (0.7) c 67.0 1927 (41) c 6.4 18 a 4.7 25.4 (1.6) d 84.2
WP 4.0 20.8 (0.9) a 15.2 1700 (70) a 8.2 32 a 4.6 16.2 (1.5) b 77.6
7.5 24.3 (1.7) b 34.3 1678 (73) a 6.9 52 b 3.8 20.6 (0.1) c 84.5
15.0 33.6 (1.8) c 85.4  1742 (108) a 5.3 95 c 5.3 28.3 (0.4) d 84.3

† Values in parentheses are standard deviations values followed by the same letter are not significantly different at P=0.05. Abbreviations: SW-switchgrass; ADF- animal digested fiber; SB- softwood bark; WP- wood pellets.

Supplemental Table 3.  Total C, C-mineralization and pool sizes of the active (Ca), slow (Cs) and resistant (Cr) C pools for Palouse silt loam soil amended with four rates of biochar pyrolized from each of four feedstocks.

C increase C                  Mineralized Active Pool Slow Pool                             Resistant Pool
Soil/Biochar CharRate Total C fromnative CumulativeCO2-C C-min /TC Ca Cs Cr Cr/TC
g kg-1 g kg-1 % mg kg-1 % mg kg-1 g kg-1 g kg-1 %
Palouse 0 23.6 (0.4) a 0 1958 (32) ab 8.3 203 a 6.5 16.9 (0.1) a 71.3
 SG 4.0 26.0 (0.2) b 10.3 1895 (55) a 7.3 159 b 5.6 20.6 (0.6) a 79.0
7.5 28.3 (0.7) c 19.8 2017 (70) ab 7.1 152 b 6.3 21.9 (1.0) a 77.3
15.0 33.4 (0.5) d 41.3 2051 (67) b 6.2 178 ab 5.5 27.8 (5.4) a 83.3
ADF 4.0 25.5 (0.3) b 8.1 2057 (47) b 8.1 176 ab 5.7 19.6 (1.9) ab 76.5
7.5 29.4 (0.7) c 24.6 2020 (253) ab 6.9 160 b 5.3 23.9(2.3) b 81.5
15.0 37.9 (0.7) e 60.7 2314 (49) c 6.1 140 b 6.5 31.3(1.6) c 82.5
SB 4.0  30.2 (1.0) c 27.9 1957 (51) a 6.5 200 a 6.9 23.0 (1.3) b 76.3
7.5  32.1 (1.7) c 35.9 2143 (45) b 6.7 182 a 7.8 24.1 (1.4) b 75.0
15.0 36.9 (0.9) e 56.4 2102 (97) ab 5.8 196 a 7.3 29.4 (2.3) c 79.6
WP 4.0   27.0 (1.5) c 14.5 2020 (30) bc 7.5 201 a 5.7 21.1 (1.1) b 78.0
7.5   35.6 (1.3) e 50.9 1940 (32) a 5.4 195 a 6.6 28.8 (2.0) c 80.9
15.0 38.3 (1.0) f 62.1 2062 (57) c 5.3 172 ab 7.3 30.8 (2.1) c 80.6

† Values in parentheses are standard deviations values followed by the same letter are not significantly different at P=0.05. Abbreviations: SW-switchgrass; ADF- animal digested fiber; SB- softwood bark; WP- wood pellets.
Supplemental Table 4.  Total C, C-mineralization and pool sizes of the active (Ca), slow (Cs) and resistant (Cr) C pools for Thatuna silt loam soil amended with four rates of biochar pyrolized from each of four feedstocks.

C increase C                    Mineralized Active Pool Slow Pool                             Resistant Pool
Soil/Biochar CharRate Total C fromnative CumulativeCO2-C C-min /TC Ca Cs Cr Cr/TC
g kg-1 g kg-1 % mg kg-1 % mg kg-1 g kg-1 g kg-1 %
Thatuna 0 26.7 (0.4) a 0 2264 (57) a 8.4 388 a 7.9 18.6 (1.4) a 70.6
 SG 4.0 28.4 (0.4) b 5.7 2482 (67) b 8.7 333 a 6.3 21.8 (2.1) ab 76.8
7.5 31.0 (0.5) c 15.4 2317 (67) a 7.5 370 a 5.4 25.3 (1.5) b 81.5
15.0 37.6 (1.5) d 33.9 2332 (32) a 6.2 314 b 7.9 29.5 (2.3) c 78.3
ADF 4.0 31.0 (0.6) c 15.4 2749 (37) b 8.9 293 b 7.4 23.4 (0.8)  b 75.4
7.5 31.8 (0.7) c 18.2 2682 (65) b 8.4 296 b 6.0 25.5(1.7) b 80.2
15.0 36.8 (0.3) d 37.0 3013 (70) c 8.2 253 b 8.4 28.2 (0.2) c 76.5
SB 4.0 28.3 (1.1) b 5.3 2156 (120) a 7.6 386 a 8.4 19.5 (0.8)  a 68.7
7.5 31.9 (1.6) c 18.5 2590 (71) b 8.1 266 b 7.3 24.3(1.8) b 76.4
15.0 34.9 (2.8) cd 29.7 2787 (85) c 8.0 297 b 8.3 26.3 (1.8) bc 75.5
WP 4.0 28.4 (0.6) b 5.6 2064 (81) b 7.3 402 c 7.5 20.5 (1.4) a 72.1
7.5 30.3(0.6) c 12.6 2214 (32) a 7.3 294 b 7.4 22.6(1.9) b 74.5
15.0 38.7(1.7) d 43.9 2054 (206) ab 5.3 423 c 6.4 32. (1.1) c 83.5

† Values in parentheses are standard deviations values followed by the same letter are not significantly different at P=0.05. Abbreviations: SW-switchgrass; ADF- animal digested fiber; SB- softwood bark; WP- wood pellets.

 

Supplemental Table 10. Total C, C-mineralization and pool sizes of the active (Ca), slow (Cs) and resistant (Cr) C pools for Hale silt loam soil amended with four rates of biochar pyrolized from each of four feedstocks.

C increase C            Mineralization Active Pool Slow Pool                             Resistant Pool
Soil/Biochar CharRate Total C fromnative CumulativeCO2-C C-min /TC Ca Cs Cr Cr/TC
g kg-1 g kg-1 % mg kg-1 % mg kg-1 g kg-1 g kg-1 %
Hale 0 39.8 (0.9) a 0 1844 (64) a 4.6 418 a 10.0 29.4 (0.9) a 73.2
 SG 4.0 42.2 (0.4) b 5.8 1866 (79) a 4.4 430 a 11.4 30.4 (0.1) a 71.9
7.5 44.1 (0.9) c 10.5 1844 (66) a 4.2 399 a 9.5 34.2 (1.1) b 77.6
15.0 48.6 (1.5) d 21.8 1686 (45) b 3.5 449 b 8.2 40.0 (0.9) c 82.3
ADF 4.0 42.6 (1.4) bc 6.8 1830 (81) a 4.3 399 a 11.1 31.2 (0.8)  a 73.1
7.5 44.7 (1.9) bc 12.0 1743 (66) a 3.9 462 b 10.0 34.2 (1.3) b 76.5
15.0 49.9 (1.7) d 25.0 1863 (106) a 3.7 550 c 9.3 39.9 (1.2) c 80.2
SB 4.0 42.5 (0.5) b 6.4 1873 (94) a 4.4 467 b 10.9 31.0 (1.1) a 73.1
7.5 43.5 (1.1) bc 9.1 1795 (57) a 4.1 397 a 11.0 32.1 (0.7  b 73.8
15.0 50.4(3.1) d 26.2 1845 (53) a 3.7 363 a 8.1 41.9 (1.1) c 83.3
WP 4.0 43.3 (0.5) bc 8.6 1854 (70) a 4.3 486 b 10.5 34.1 (3.1) ab 78.7
7.5 45.7 (1.3) c 14.5 1774 (30) a 3.9 343 d 9.8 35.5 (2.6) b 78.8
15.0 51.5 (0.8) d 28.9 1783 (81) a 3.5 314 d 9.3 41.9(1.4) c 81.4

† Values in parentheses are standard deviations valuesfollowed by the same letter are not significantly different at P=0.05. Abbreviations: SW-switchgrass; ADF- animal digested fiber; SB- softwood bark; WP- wood pellets.

Chapter Three

The Ability of biochar produced from Anearobic fiber to reduce phosphorus levels in dairy lagoons and its characterization

 (Prepared for submission to Journal of Environmental Quality)

ABSTRACT

Rising production costs and environmental concerns over phosphorus fertilizer and dairy nutrient management have prompted investigations of phosphorus recovery from anaerobic digesters.  We evaluated the use of biochar produced from anaerobic digested fiber (APF) for sequestering phosphorus from dairy lagoons.  The APF was collected from a Plugged Flow (GHD™) digester; air dried to < 8% water content, pelletized under pressure (2250 kg/cm2) and stored dry.  To produce biochar a barrel retort slow pyrolysis unit was constructed.  Phosphorus recovery and reduction potential of biochar in ADE was measured with small scale filter systems (1:25 biochar to ADE). The analysis showed average reductions of  381 mg P L-1 and 4 g L-1 (75%) of fiber. There was an increase of total, Olsen, and water extractable P on the char after 15d at 1.9 g      kg1, 763, and 914 mg kg-1, respectfully. This accounted for a 32% recovery of P from the orginial lagoon. The recovered P on the biochar was then analyzed using 31-P NMR for P speciation.  NMR confirmed the recovery of inorganic P on the biochar after filtration and the presence Ca-P in the solid form of the biochar.   Our research indicates biochar could be beneficial component to P reduction in the dairy system.


INTRODUCTION

Dairies in the state of Washington produce over 1.5 million tons of manure annually accounting for 19 Mg of phosphorous (P) that is concentrated in storage lagoons prior to application to agricultural lands (Frear et al., 2005). Annual land applications have led to concerns of environmental degradation caused by NO3 leaching and P runoff. These applications typically deliver > 600 kg N ha-1 y-1 and 160 kg P ha-1 y-1 exceeding crops needs (Hart et al., 1997). The generational manuring of lands adjacent to dairies has led to excess P accumulation causing serious health and environmental problems such as eutrophication (Sharpley et al. 1995; Sharpley 2000; Pierzynski, et al., 2000). As a result of environmental degradation, regulations and standards have been developed limiting the amount of nutrients from dairy manure which can be land applied annually (Clean Water Act).  The problem of excess soil P in conjunction with concerns over limited future supplies of P has created a situation where low cost P recovery systems are needed. One method being utilized to manage on-farm nutrients generated by concentrated animal feeding operations (CAFO) throughout the United States is the use of anaerobic digesters (AD) (www.epa.gov/AgStar).  Anaerobic digesters reduce odor, solids and greenhouse gases but maintain the nutrient content of the effluent that is stored in lagoons. (Kaparaju and Rintala, 2010; Hobson and Feilden, 1982; Rico et al., 2007; Gungor and Karthikeyan, 2008; MacConnell and Collins, 2009). Ideally the excess P and N within the effluent stream could be recovered as a fertilizer and exported off farm rather than applied on farm.

As the demand for phosphorus increases over the next 50 years in response to world population growth supplies of rock phosphate will be depleted (Cordell et al, 2009).Recovery of P from the AD effluent stream will improve  nutrient management and mitigate the environment concerns surrounding the dairy industry.  The prospect of P reclamation and storage in plant available form can potentially lead to sustainability of the global phosphorus cycle.

The application of P recovery technology has been based on the use of precipitation or adsorption (Bashan et al., 2004).  Bashan et al. (2004) described the use of  iron, aluminum, magnesium and calcite products from industrial wastes such as a slag, or oxides from steel production in waste water P removal. Dou et al. (2003) found alum, combustion fly ash, fly ash, and a desulfurization by-production all reduced P in swine (Sus domesticus), chicken (Gallus gallus domesticus) and dairy cattle (Bos indicus )manure.  In the dairy system there has been a revitalization in the use of struvite (magnesium ammonium phosphate hexahydrate, NH4MgPO4-6H20) to recover P from various waste streams (Westerman et al, 2009; Massey et al., 2007; Munch and Barr, 2001; Schuiling and Andrade,1999). The implementation of engineered cone-shaped, fluidized bed systems allowed for greater control over the struvite formation process proved successful in swine and dairy systems (Bowers et al., 2007; Bowers and Westerman, 2005).  Activated carbon (AC) from various sources has also been used for P removal.  The use of AC from multiple sources of waste streams has been a proven method for containment removal, however activation remains expensive.  This expense has kept it out of the waste management programs of the dairy industry.

An alternative potentially cost effective approach for P removal in dairy effluent besides AC and struvite may be the by-product of pyrolysis of the fiber separated from the dairy manure.  The thermo-chemical conversion (pyrolysis) of plant biomass is being studied globally for its potential to produce second-generation liquid biofuels and the co-product biochar (Sohi et al., 2010; other Citations). A carbon-rich product remaining after biomass (biochar) has been heated at temperatures ranging from 350°C to 600°C in an oxygen free environment (Amonette and Joseph, 2009; Graham et al., 1984; Czernik et al., 1994; Bridgwater et al., 1999; Mohan et al., 2006; Huber and Corma, 2007).  High cation exchange capacity (CEC; 40 to 80 meq 100 g-1), high surface area (51 to 900 m2 g­1), and affinity for micro- and macro- plant nutrients are among the beneficial characteristics of biochar (Lehmann, 2007; Laird, 2008; Gaunt and Lehmann, 2008; Cheng et al., 2008; Novak et al., 2009; Lehmann and Joseph, 2009; Roberts et al., 2010; Streubel et al., 2011).  These beneficial characteristics of biochar have been linked to the highly organized C-ring structure of biochar (Mykola et al., 2008; Keiluweit et al, 2010).  The characteristics of biochar structure and their influence on soil may also help in P recovery from dairy effluent.

Biochar has previously been produced using animal manures as a feedstock (Ro et al., 2009; Szogi and Vanotti, 2009; Lima and Marshall, 2005 ; Cao and Harris, 2010; Shinogi and Kanri, 2003;Chan et al., 2008) Streubel et al. (2011) evaluated the fiber (ADF) from an anaerobic digester as a source of biochar as a soil amendment and found it raised pH, and added C in four soils while showing mixed results in its able to increase water holding capacity.  Chen (2010) found that biochar derived from orange (Citrus Sinensis (L.) peels enhanced with iron effectively removed phosphate and organic containments from waste water while granular activated carbon from poultry manure removed 1.92 mmol Cu++ g-1 of carbon (Lima and Marshall, 2005).   Szogi and Vanotti (2009) include biochar production from poultry manure for remediation as a potential approach to improve nutrient management in that industry.   Properties of dairy-manure derived biochar were found to be favorable for the absorption of Pb (192 mg     L-1or 93%) and Atrazine (3 mg L-1 or 77%) from a liquid solution (Cao and Harris, 2010). Although the use of biochar has been studied recently there remains a lack of literature on the use of ADF biochar as a P removal media in anaerobically digested effluent (ADE).

The objective of our research was to evaluate a low cost methodology for P recovery where no further chemical enhancement of biochar was necessary by 1) determining the biochar ability to remove P from ADE; 2) determining the concentration and availability of P on the biochar after being used as filter media and 3) characterizing the P form on the biochar.  Our main hypothesis for this research is that ADF biochar will remove P from the ADE and retain the P on the biochar because of its physical structure and chemical bonding.

Materials and Method

Biochar Production

Animal-digested fiber (ADF) separated from a GHD Inc. (Chilton, WI) Plugged Flow anaerobic digester located on a dairy in Outlook, WA was used as the biochar feedstock. The ADF was air dried to less the 80g kg-1 moisture and commercially pelletized  (Mid-Valley Milling Inc. Prosser, WA) at 207 MPa into 5 mm diameter pellets.  Biochar was produced from the ADF pellets at a pyrolysis temperature of 500oC for four hours using a batch barrel retort manufactured by Marv’s Reliable Maintenance, Bellingham, WA. Pyrolysis temperature of the retort was monitored with XCIB-K-1-6 thermocouples (Omega Incorporated, Stanford, CT) mounted inside the retort and connected to a Campbell Scientific X10 Data Logger programmed to continuously (2 minute) collect data for temperature verification (Figure 1). Upon batch completion the retort was allowed to air cool then biochar was removed and stored dry.

Laboratory Analysis of Biochar P Retention

Four 10 g sub-samples of biochar were incubated with 50 ml of ADE (40 mg L-1), a K2HPO4 (38 mg L1) standard and di-ionized water (0.02 mg L-1) control.  Samples were put under oscillation in 50 ml centrifuge tubes at 25 oC for 48 h.  After shaking samples were decanted and air-dried at 21oC.  After drying each sample was then extracted to determine Olsen P concentration (Kuo et al. 1995).  Briefly, 2 g subsamples of biochar were placed in 40 ml of 0.5 M NaHCO3 at a pH of 8.5, shaken for 30 minutes and filtered through a Whatman 42 filter.  Each extract was stored at 5oC until P analysis on a multiple sample spectrophotometer  (EasyChem Analyzer, Systea Scientific, Oak Brook, IL)  using the ascorbic acid colorimetric method. This experiment was conducted three times with 4 replicates for each treatment for a total of 12 samples.

Experimental Lagoon Design

Digested effluent was collected immediately within the first settling chamber exiting the GHD Plugged Flow digester and before transfer to the dairy storage lagoon. The effluent was collected with a 5.5 hp Buffalo WP20 water pump (Buffalo Tools, St. Louis, Missouri) and stored in a closed 946 L plastic tote for no longer than 2 h prior to study initiation.  Three 560-L water tanks (High Country Plastics, Caldwell, ID) were filled with 378 L of the effluent. Each experimental lagoon contained a Laguna Max-Flo 2000 submersible pump (Rolf C. Hagen Corp. Mansfield, MA) connected to a Tetra Pond Bio-Active Pressurized Filter BP 4000 (Tetra Werke Co, Melle, Germany) containing 15 kg of ADF biochar (Figure 2).  Effluent was cycled continuously through the filter system at a rate of 2400 Lh-1 and sampled on days 0, 5, 10, and 15 for P concentration.  Effluent samples (1-L) were collected from the center of each experimental lagoon during active filtration.  Biochar samples (100 g) were collected from the center of each filtration unit after shutting down the filers. After sub-sampling each system was restarted.  Samples were stored at 5oC until analyzed.   Lagoon effluent was analyzed at a commercial lab (Soiltest Consultants, Moses Lake, WA) using nitric acid digestion (Gavlak et al., 1994). Briefly nitric acid is added to effluent and air dried for 24 hrs then diluted with 40ml of DI water and totals determined by inductively coupled plasma atomic emission spectroscopy (ICP; DV3300, PerkinElmer, Waltham, MA) according to Isaac and Johnson (1998).  Macro- and micro-elemental composition (P, K, S, B, Ca, Cu, Fe, K, Mg, Mn, and Zn) of biochars before and after filtration were also determined by ICP.  Each biochar sub sample was dried at 45oC for 48 hrs and analyzed for total, Olsen, and water extractable P (WXP).  Briefly, water extractable P was collected by shaking 5 g of incubated biochar-soil in 50 ml of deionized water for 1 hour.  Immediately upon shaking samples were centrifuged for 30 minutes at 335 Hz.  Filtrate was then decanted and filtered using a Whatman 42 filter. The pH of lagoon effluent was determined using a pH meter (Model 445, Corning Inc., Corning, NY) directly placed in 30 mL of the effluent subsample.

31P NMR Analysis.

Initial biochar and biochar samples collected from the lagoons were prepared for 31P NMR analysis using three extraction methods:  a water extraction (WXP), 1 M HCL and 0.25 M NaOH-EDTA extraction. For the HCL extraction 2 g of sample were added to 40 ml of 1 M HCL.  In the NaOH-EDTA extraction 2 g of sample were added to 40 ml of NaOH –EDTA solution comprised of 0.25 M NaOH mixed with 0.05 M EDTA at a 1:1 ratio.  The WXP used 5g of biochar to 50ml of H20.  The WXP and HCL extractions were shaken for 30 minutes and centrifuged at 4oC for 30 minutes at 166 Hz. The NaOH-EDTA extractions were shaken for 16 hours then centrifuged at 166 Hz for 30 minutes.  All sample supernatants were frozen for later analysis (He et al. 2003).  Three samples were extracted for each method and biochar then a composite of the three extracts were used for analysis.  Each extract was analyzed by solution state 31P NMR spectra using a Varian Unity Plus 500 MHz spectrometer (Walnut Creek, CA) equipped with a 10-mm HX probe tuned to 202.31 MHz. The pulse sequence used a single 45o pulse, 0.64 second acquisition time, and 1.200 second recycle delay.  Temperature was regulated between 20 and 22o C, and 28,000 to 30,000 scans were acquired per sample.  The phosphorus chemical shift scale was standardized to 0 ppm, using an external 85% phosphoric acid standard.  Phosphorus compounds were identified by their chemical shifts (ppm) and compared to literature databases for P speciation.

Samples of biochar prior to and after pond filtration and biochar samples after each chemical extraction described above were analyzed with solid-state 31P NMR.  The spectra were collected with a Varian Unity Plus 500 MHz spectrometer operating at 202.311 MHz equipped with a 5 mm HX MAS probe with zirconia rotors. The spinning rate was 10 kHz, and the pulse sequence used a 45 o pulse followed by a cycle delay of 30 seconds.  A total of 2400 to 30,000 scans were collected to acquire sufficient signals to noise ratios.

FIB/SEM Analysis

The detailed structures of selected biochars were examined using a FEI Helios Nanolab (FEI, Hillsboro, OR) dual-beam focused ion beam/scanning electron microscopy (FIB/SEM) microscope. Intact and crushed microscopic structures of biochars were observed in cross sections (300um – 2mm). To prepare the sample, a piece of selected intact or crushed biochar was mounted on a 1 mm diameter stub. The samples were carbon coated for SEM observation to eliminate reflectance from Fe content in the biochar.

Statistical Analyses 

Data were tested for normality by Shapiro-Wilk’s W test and for heterogeneity of variance by Levene’s test using SAS (version 9.2, SAS Institute, Cary, NC). Associations between day, nutrient analysis, pH, WXP, Olsen P, and lagoon P concentration were assessed using PROC MIXED, PROC GLM and ANOVA with pairwise comparisons by Tukey at P < 0.05 (version 9.2, SAS Institute, Cary, NC).

Results and Discussion

Feedstock and Biochar C, N, and S

The C concentration of the feedstock and pellet prior to pyrolysis was 393.1 and 390.3 g kg-1 (Table 1).  The N and S concentrations of the fiber were 25.5 and 7.4 g kg-1 while the pelletized fiber had lower concentrations of 21.4 and 5.9 g kg-1.  The mineral and organic content of the feedstock remain unchanged during pelletization with 80 % of the biomass remaining as organic material.  Pyrolysis of the ADF feedstock concentrated the ash content of the biochar increasing from 200 to 425 g kg-1 as a result of C loss (Table 1).  These results are consistent with Streubel et al. (2011) who reported a similar increase in ash of herbaceous feedstocks which included ADF.  Similar findings were also reported by Cherney et al. (1988) and are found in the Phyllis Database for Biomass Waste (2010).  The high organic content of the fiber is due to the composition of hemicellulose, cellulose and lignin found in herbaceous grasses which are traditionally fed to dairy cows for milk production.    Table 2 shows the elemental composition of the fiber and biochar prior to and after treatment with digested dairy effluent. In conjunction to the P the Ca, Mg and Fe concentrations were 29, 11 and 8 g kg-1 in the biochar prior to filtration and 35, 12, and 10 g kg-1 in the biochar after filtration of effluent over a 15d period.  These higher concentrations are a direct reflection of the manure feedstock. The concentration of the remaining elements showed little change from the untreated biochar (BI) are likely strongly associated within the C matrix.    In a review of several horticultural based feedstocks in the literature Schiemenz and Eichler-Lobermann (2010) found concentrations ranging from 1 g kg-1 in bagasse ash to 20.1 g kg-1 in straw ash.  It is the presence of these elements in higher concentration makes struvite production from dairy manure limited without considerable effort in modifying the process and expense (Bowers and Westerman, 2005).

Affect of Biochar on AD Effluent pH 

In a preliminary laboratory experiment the addition of biochar (pH 9.3) to AD effluent (1:5) increased pH  from 7.6 to 8.6 within 5 hours after addition, increasing to a final pH of 8.8 after 90 hours (Figure 2). The same change in pH was observed in the experimental lagoons (Figure 3.0). However, the change in pH took longer to increase in the experimental lagoons then laboratory and was likely due to the difference in the ratio of biochar to effluent in the laboratory (1:5) and experimental lagoons (1:25). In this aquatic system the surface chemistry of the biochar is most likely influencing the pH change.

The effect of biochar on altering the pH of dairy effluent has not been previously reported in the literature; however a review of carbon chemistry and the characteristics of activated carbon may explain the results. The surface chemistry of biochar is complex, being dependent upon feedstock characteristics and the temperature and method of carbonization (slow vs. fast pyrolysis) (Lehmann and Joseph, 2009; Amonette and Joseph, 2009; Baldock and Smernik, 2002; Antal and Grǿnli, 2003; Krull et al., 2009; Chan and Xu, 2009). The mineral composition of the biochar creates a heterogeneous mixture of both electron donors and electron acceptors that co-exist on the surfaces of biochar (Downie et al., 2009).  Farmer et al. (1996) found that when activated charcoal (AC) was used in waste water treatment pH increased from 7.4 to 10.5.  They suggested that in the presence of water H+ ions hydroxides of Ca K and Mg on the surface of AC could neutralize acidity    A similar explanination change in pH was given for three ACs used for removing reactive dyes in the textile industry (Al-Degs et al., 2000). The donor-acceptor chemistry was found to come for the oxygen groups interacting with carbonyl and phenol aromatic rings in AC (Franz et al., 2000).  It has also been suggested that if the carbon source has less acidic functional groups the carbon will preferentially absorb water through hydrogen bonding therefore acting like a base for the solution (Li et al., 2002). Nevskaia et al. (2004) found that interactions with AC surface chemistry and water was also dependent on the presence of oxygen surface groups.

Phosphorus Pool Distribution in the Experimental Lagoons

Phosphorous concentrations among experimental lagoons prior to the start of the filtration through biochar ranged from 410 to 610 mg P L-1 (Table 3). Total P concentrations included the fiber component of the effluent. After the first 5 days of filtration an average of 60% of the P was removed from the lagoons (Table 3). At the end of 15d, P concentrations in the lagoons decreased 70% from an average of 546 mg P L-1 to 165 mg P L-1 (Table 3). The average reduction in P among lagoons was 381 mg P L-1, suggesting the biochar removed 144 g of the initial 206 g P from the lagoons (Table 3). A control lagoon without biochar in the filter system showed was set up for 5 d no change in P concentrations was recorded (data not shown).

The initial solid fraction (fiber) content of the effluent ranged from 3.6 to 6.7 g L-1 (1.4 – 2.5 kg-1 lagoon) among lagoons (Table 3). The fiber originated from the undigested plant materials present in the effluent following anaerobic digestion of the manure. The biochar after filtration had a matrix of coatings that included some identifiable fiber materials surrounding the biochar pellet.  These coatings account for 5 – 8 % of the total weight of the Bd15 samples (data not shown).  The coatings can be seen in the SEM images in Figure 5. The mass of fiber suspended in the effluent was reduced an average of 75% (4 g L-1) by filtration through the biochar over the 15d of the study (Table 3).  Total fiber extracted by the biochar averaged 2 kg lagoon-1 (133 g fiber kg-1 biochar).

The adsorption of phosphate ions will depend on the concentration and accessibility of cations found in the ash (Lu, et al., 2009; Agyei et al., 2000; 2002; Namasivayam and Sangeetha 2004; Oguz, 2005; Zhang, 2007; Xue et al., 2009). The use of fly ash has been studied for its potential to remove phosphate compounds from waste water. Only a portion of the research has been devoted to understanding how different metal ions in the ash affect removal from solutions. The application of this information to bio-chars for the removal of relatively low concentration phosphates is of interest, because of the desire to develop an adsorbent capable of removing nutrients from waste streams.

Other research has focused on inducing the precipitation of struvite (magnesium ammonium phosphate hexahydrate, NH4MgPO4-6H20) for recovering P from various waste streams (Westerman et al, 2009; Massey et al., 2007).  Schuiling and Andrade (1999) developed an approach to recover struvite from calf and chicken manure.  Australian scientists reduced P concentrations by 61 mg/L from waste water treatment sludge using magnesium hydroxide to force the precipitation (Munch and Barr, 2001).  Tamarind (Tamarindus indica Linn)  nut shell AC was found to successfully remove 95% of the P from a standard solution through adsorption (Bhargava and Sheldarkar, 1993). Powdered AC from hardwood was found to remove up to 2 mg L1 of industrial waste water effluent P (Meidl, 1997).  Along with almond (Prunus amygdalus Batsch) and pecan (Carya illinoinensis) shells AC shells being used as AC sources for P removal (Toles et al., 1997) poultry manure AC has been used to remove mercury from coal fire plants (Klasson et al., 2010;Yaji et al., 2008). Effects of metal ion concentrations in fly ash have been the subject of significant research (Lu, et al., 2009; Agyei et al., 2000; 2002; Namasivayam and Sangeetha 2004; Oguz, 2005; Zhang, 2007; Xue et al., 2009).   The adsorption of phosphate ions will depend upon the concentration and accessibility of cations found in the ash which is important for maximizing recovery.

Phosphorus Mass balance following filtration

Total P concentration of the biochar prior to effluent application was 9.2 g P kg-1, with 26 and 200 mg P kg-1 water and NaHCO3 extractable-P, respectively. After 15d of effluent filtration the treated biochar had a total P content of 11.1 g P kg-1, with 784 and 963 mg P kg-1 released by the water and NaHCO3 extractions, respectively. The sorption of effluent P by the biochar represented a 29 and 5 fold increase in WXP and NaHCO3-extractable P above the untreated biochar, respectively. Figure 6 shows the change in WXP and NaHCO3-extractable P over the 15d of filtration, indicating that 60 mg P kg-1 d-1 of WXP and 62 mg P kg-1 d-1 of NaHCO3-extractable P was adsorbed by the biochar over the 15d treatment period. Over the 15d period WXP increased on the biochar from 26.1 mg kg-1 of biochar to 940 mg kg-1 of biochar an increase of 914 mg kg-1 (Figure 6.0).

At each interval of sampling sodium bicarbonate extractable P increased linearly over the 15d period.  The increase reached 800 mg kg-1 at 15d (Figure 6).  The NaHCO3 (Olsen P)  chemical extraction  method was developed for determining the labile P pool which is mineralized over the growing season and in turn becomes available to the plant. The Olsen method is best suited for alkaline, calcareous conditions with a pH above 7.2 (Marx et al., 1999; Elrashidi, 2008).  The NaHCO3 extraction appears to represent the same pool of adsorbed P as determined by the water extraction (Figure 6). Each of these extractions was conducted on independent samples and not performed sequentially. Subtracting the background NaHCO3 extractable-P of the untreated biochar from the d15 coated biochar yields 763 mg P kg-1, similar to the WXP. This pool of soluble P (both WXP and NaHCO3 extractable-P) was assumed to be present in the pores of the biochar and spaces between biochar pellets and held by capillary and matric potential forces upon gravitational draining after removal of the biochar from the filter. This pool of P was fixed/adsorbed to the surface of the biochar during the drying process. Incorporation of a sequential extraction in the analysis will be necessary to clarify these results. The biochar was also coated with 133 g fiber kg-1 biochar which had a concentration of 8.5 g P kg-1 fiber that contributed an additional 1.13 g P kg-1 to the biochar matrix.

The mass balance of P in the lagoon/filter system showed that on average; 144 g P (381 mg P L-1 *378 L * 10-3 g P mg-1 P) was removed from the lagoon effluent with 29 g P (1.9 g P kg-1 biochar *15 kg char) adsorbed to the biochar in the filter and an additional 17 g P (1.13 g P kg-1 *15 kg char) was present in the fiber coating the biochar. These two pools sum to 46 g P associated with the biochar upon drying and result in a recovery of 32% of the original P in the lagoon by the biochar.  The remaining 68% of the P (98 g P) was assumed to be weakly associated with the biochar/fiber matrix and lost in drainage waters when the biochar was removed from the filter. This “lost” P was not accounted for or represented by the P concentration in the lagoons at d15. An analysis of the drainage waters from the biochar in the water of the filter would clarify this assumption. A short term laboratory experiment was conducted in an attempt to verify these assumptions. A 1 g sample of BI was placed in a 50 ml centrifuge tube with 26 ml of ADE and oscillated for 5 days.  Upon completion liquid was either dried at 100oC or decanted and then all samples were dried.  Samples showed an increase of Olsen P on both treatments, however the dried down samples had 1984 mg kg-1 greater amount of Olsen P than the decanted samples.

Nuclear Magnetic Resonance Spectroscopy  

Our results demonstrate the first attempt at identifying the forms of P recovered by biochar from an anaerobic digested effluent.  The water soluble or readily available forms of P were extracted using the WXP method, while the stable fractions were extracted using HCl and NaOH-EDTA.   In the NMR analysis major peaks where determined for the liquid and solid fractions for the untreated and d15 effluent coated biochar that resulted in differences in P speciation signals.  The P fractions present in the solutes after extractions of the untreated biochar with WXP, HCl and NaOH-EDTA displayed peaks at 0.359, 2.9 and 5.210 ppm, respectfully (Figure 7).  The three extractions on the d15 biochar had peaks located at 0.459. 3.120 and 5.25 ppm, respectfully (Figure 7).  For the d15 biochar the EDTA extraction also showed a peak in the 0 ppm area of the chemical shift spectrum.  We assigned this peak to the presence of DNA.  This peak can indicate the presence of coatings on the biochar and microbial biomass.  This is likely as the coatings are fibrous plant material and come from a microbe rich environment.  The peak locations were similar for between the untreated biochar and the d15 biochar, however the extracted P fractions were different.  The total P extracted from the biochar in WXP, HCl, and EDTA was 0.3, 1.9, and 1.1 g kg-1, respectively (Table 4).  For the  d15 biochar extracted P fractions were 10 and 2 times higher for WXP and NaOH-EDTA; 0.3 and 1.8 g kg-1, respectively (Table 3). The HCl extraction removed twice as much P and Ca from the untreated biochar.  We suspect, although did not confirm, that the concentration of HCl was not high enough to release P from the coated d15 biochar.  We suggest that either increasing the concentration of the HCl or number of extractions would verify this hypothesis.

The liquid NMR results show that all the dominate species of phosphorous was inorganic orthophosphate with minor peaks represented by DNA, and RNA.  The primary peaks in each of the extractions found on the spectra are most likely an effect of the pH change in the extraction liquid rather than a large shift in P species.  Our results were consistent with other liquid NMR of manures with the common peak of inorganic orthophosphate, however in our samples there was no detection of some P compounds such as mono-and-diesters, and polyphosphate.  This absence could be an effect of the anaerobic digestion.  (Turner and Leytem, 2004; Turner et al., 2003; He et al., 2009; He et al., 2007).  The absence of other visible peaks could be because the samples were from anaerobic digested effluent and dried. McDowell et al. (2007) suggest once manures are dried the P species transform most likely to the orthophosphate form.  The results is surprising that no phospholipid peaks were determined in the 15 day coated biochar as the coats are the remaining undigested fiber from forages where are still visible on the biochar (Figure 8).  The DNA found on the 15 day biochar is mostly likely from the plant tissue and bacterial populations in the AD effluent. The intensity of the peaks in the 15 day coated biochar are consistent with the increase of water and Olsen P extractable P.  The difference is the HCl which is reversed.  The HCl extracted the high amount of Ca and this could be a direct result of the pH of 3.0 nature of the extracting HCl liquid as Ca dissolves in acidic solutions.  Turner and Leytem (2004) suggest the use of HCl extraction appears to hydrolyze and precipitate Ca because of the pH value of the extract.

In the solid state NMR the peaks range from -1.202 to 1.808 for extractions and controls on the spectrum shift but show a broad hump (Figure 10.0).  These visible shifts in the solid state form are not significantly different from one another to make specific P form spectral assignments, however the elemental analysis of the solid material after extractions (Table 4. shows shifts in the amount of P in the solid forms after the extractions.  The presence of spinning side bands (SSB) at 50 and -50 show the presence of metal compounds in the solid structure (He et al., 2009).  In conjunction with the elemental analysis and others research we can determine that the metal compounds are likely Ca and Mg with Ca being dominate.  In the WXP extracted form the peak of -1.4 makes CaHPO4 where the other extractions cover other forms of metal HPO4 and  H2PO4 bonds (He et al. 2007).  Our results are consistent with the literature.  The lack of major shifts between the biochar and 15 day coated biochar is likely because of the fixed structure of the carbon base of biochar.  The carbonization process leaves carbon rings fixed with a random mixing of other metals, elements and functional groups.  The biochar functional groups are a direct correlation between the elements found in the feedstock (Downie et al., 2009) As the original material was cow manure it is logical the Ca and Mg would be present in the solid form.  This is confirmed by the elemental analysis (Table 4).

 Conclusion

Results from this research show biochar can be manufactured  from ADF as a material for the recovery of P from the effluent stream inferring its potential use a remediation tool.  We recovered 30 % of the P from the effluent over a 15 day period.  The P recovery is dominated by the water extractable forms dried down onto the biochar.  The use of WXP and Olsen P extractions showed significant increases of plant available P on the biochar.  The liquid and solid state 31-P NMR confirmed the inorganic species presence on the biochar while the dominate presence of Ca-P complexes in the solid.  Our NMR studies did not report a quantitatively reading for P concentration, however gave us qualitative look at differences confirming the results from the extractions.  While biochar has been studied for soil impacts this is the first data representing biochar produced from ADF and used to recover P from AD effluent.  Although recovery was less than anticipated we have shown the concept is sound, as even at a 30 % reduction and recovery this represents a benefit to both the environment and dairymen.

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Table 1. Selected characteristics of the four stages in the material process.

Characteristics

C

N

S

C:N

C:S

Ash

Source

—————— g kg-1 ——————

g kg-1

ADF

  393.1 (0.4)

25.5 (0.04)

7.4 (0.05)

15

53

194.7 (20)

ADFP

390.3 (0.1)

21.4 (0.01)

5.9 (0.003)

18

67

209.0 (40)

BI

483.5 (1.2)

21.7 (0.07)

5.5 (0.03)

22

89

431.6 (1.2)

Bd15

461.9 (0.3)

22.0 (0.07)

5.7 (0.03)

21

82

424.5 (20)

ADF-anaerobic digested fiber; ADP- anaerobic digested fiber Pellet; B- anaerobic digest fiber pellet biochar; BD15- anaerobic digest fiber pellet biochar 15 days in contact with effluent

Std. deviation in parentheses.

Table 2. Concentrations of phosphorus and fiber in each experimental lagoon.

Phosphorus (mg L-1)

Fiber (g L-1)

Experiment

Lagoon

Day 0

Day 5

Day 10

Day 15

Total Reduction

Day 0

Day 5

Day 10

Day 15

Total Reduction

Control

610

610

610

610

0

6.7

6.7

6.7

6.7

0

1

A

410

200

160

140

270

3.6

2.4

2.0

0.9

2.7

2

A

610

241

160

182

428

6.7

4.6

1.5

1.4

5.3

B

610

170

140

140

470

6.7

3.0

2.0

0.9

5.8

C

610

173

150

140

470

6.7

2.6

2.7

0.9

5.8

Average

610 (0)

195 (40)

150 (10)

154 (24)

456 (24)

6.7 (0)

3.4 (1.1)

2.1 (0.6)

1.1 (0.3)

5.6 (0.3)

3

A

530

180

265

140

390

3.7

3.1

2.4

2.2

1.5

B

530

230

185

180

350

6.2

2.8

2.6

0.9

5.3

C

520

320

237

230

290

3.7

2.8

3.4

2.0

1.7

Average

527 (6)

243 (71)

229 (40)

183 (45)

343 (50)

4.6 (1.5)

2.9 (0.2)

2.8 (0.5)

1.7 (0.7)

2.8 (2.2)

 

 

 

 

 

 

 

 

 

 

 

All

546 (73)

216 (54)

185 (48)

165 (35)

381(81)

5.4 (1.6)

3.1 (0.7)

2.4 (0.6)

1.4 (0.6)

4.0 (2.0)

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

Difference between day 0 and day 15.  ‡Standard deviations in parentheses. Experiment 1 only one replicate; all lagoons contained 378 L of effluent with 15 kg biochar per filter at a ratio of 25:1.

Table 3. Elemental composition of manure pellet before and after pyrolysis and three periods during the 15d filtration with effluent.

Elemental Composition

Source

P

K

Ca

Mg

Zn

Mn

Cu

Fe

Na

——————————— g kg-1 biochar ——————————-

Fiberpellet

8.5 (0.6)

  9.4 (1.1)

16.0 (2)

  5.6(0.5)

0.17 (0.02)

0.20 (0.02)

0.04 (0.01)

  3.5 (0.4)

2.6 (0.2)

BI

  9.2 (0.3)

15.4 (0.4)

28.8 (4)

11.1(1.7)

0.27 (0.01)

0.40 (0.01)

0.06 (0.01)

  7.5 (0.3)

4.9 (0.2)

Bd5

  9.4 (0.2)

11.7 (1.6)

30.1 (0.8)

10.6 (0.2)

0.25 (0.01)

0.40 (0.02)

0.07 (0.01)

  7.5 (0.3)

3.6 (0.6)

Bd10

  9.7 (0.3)

10.4 (1.2)

30.8 (0.8)

10.8 (0.5)

0.26 (0.01)

0.41 (0.02)

0.07 (0.01)

  7.7 (0.3)

3.5 (0.4)

Bd15

10.7 (0.8)

14.8 (1.3)

38.2 (0.8)

12.2 (0.3)

0.35 (0.08)

0.48 (0.01)

0.10 (0.01)

10.4 (0.3)

3.5 (0.5)

Fiberpellet – prior to pyrolysis; BI – biochar after pyrolysis; Bd5, Bd10 and Bd15– biochar pellet after 5, 10, and 15 days of filtering through AD effluent. Values in parentheses are standard deviations.

Table 4. Elemental composition of liquid and solid fraction after extractions of the initial biochar and biochar after 15d of filtration through anaerobic digested effluent used in 31P NMR analyses.

Element Composition

P

K

S

Ca

Mg

Na

Zn

Mn

Cu

Fe

B

Liquid fraction

——————– g kg-1 ———————

————— mg kg-1 —————-

BI

WXP

0.03

1.80

0.08

0.01

0.02

0.35

  0.07

  0.30

0.20

0.60

0.20

HCL

1.90

4.76

0.25

7.76

1.40

1.10

16.00

40.30

0.87

80.90

2.80

NaOH-EDTA

1.11

5.00

2.79

4.37

0.32

3.44

  8.20

10.30

0.60

5.60

0.60

Bd15

WXP

0.31

3.27

0.17

0.03

0.08

0.80

  1.60

  0.12

0.18

3.30

2.10

HCL

0.80

2.04

0.08

3.02

0.68

0.37

10.60

16.20

0.27

83.40

1.10

NaOH-EDTA

1.83

7.57

1.50

4.36

0.55

3.40

12.80

15.90

1.06

9.50

3.30

Solid fraction

   ———————————————— g kg-1 ————————————————

BI

Total

10.9

19.6

4.2

38.2

11.6

4.4

0.3

0.5

0.1

9.7

0.05

WXP

10.6

18.2

3.9

38.8

12.0

4.0

0.4

0.5

0.1

10.7

0.04

HCL

8.5

14.5

3.2

30.0

10.6

3.3

0.3

0.4

0.1

9.7

0.03

NaOH-EDTA

10.5

13.7

3.7

35.8

12.2

9.6

0.3

0.5

0.1

10.3

0.04

Bd15

 

 

 

 

 

 

 

 

 

 

 

Total

11.5

19.4

4.9

35.6

12.6

4.3

0.6

0.5

0.1

9.1

   0.05

WXP

10.7

13.3

4.3

37.7

11.8

3.1

0.5

0.5

0.1

10.6

0.04

HCL

9.8

11.5

3.8

34.4

11.5

2.6

0.3

0.4

0.1

10.7

0.04

NaOH-EDTA

9.9

11.3

3.7

34.8

12.0

8.6

0.4

0.5

0.1

10.4

0.04

BI – biochar after pyrolysis; Bd15– biochar pellet after 15 days of filtering through AD effluent.

Figure 1. Barrel retort temperature recorded over the 4 hour pyrolysis period.

Figure 2. Diagram of anaerobic digested effluent biochar recovery experimental set up.

Figure 3. Change in effluent pH following the application of biochar during

a 90 h laboratory experiment. The ratio of AD effluent to biochar was 5:1.

Figure 4. Change in lagoon effluent pH after continuous flow filtration through

biochar over a 15d period.  The ratio of AD effluent to biochar was 25:1.

Figure 5. Scanning electron micrographs of biochar prior to filtration (A) and (B) biochar 15 d after used as a filter media for anaerobic digested dairy effluent.  Micrographs taken at the Environmental Molecular Sciences Laboratory, Richland, WA.

Figure 6. Change in NaCHO3 (Olsen) and water extractable P  (WXP) from biochar after continuous filtration of anaerobic digested effluent over a 15d period.

B

A

Figure 7. Liquid State NMR for biochar prior to filtration (A) and (B) biochar 15 d after used as a filter media for anaerobic digested dairy effluent.

B

A

Figure 8. 31P Solid State NMR spectra for biochar prior to filtration (A) and (B) biochar 15 d after used as a filter media for anaerobic digested dairy effluent.

APPENDIX

                  Appendix Figure 1 Batch Retort Pyrolysis unit constructed for biochar production

Appendix Figure 2 378 L Experimental lagoon set up in Prosser, WA
CHAPTER FOUR

Phosphorous, N and C availability from biochar

 amended with dairy effluent

ABSTRACT

A series of laboratory experiments were conducted to determine the effects of biochar (BI) and a biochar (BA) amended with dairy effluent on the availability of C, N and P in a Quincy sand (Xeric Torripsamment). Soils were amended with 10, 20, and 40 Mg ha-1 of each of the biochars in crushed and pellet forms and evaluated for N mineralization, C-mineralization, sodium bicarbonate extractable P (Olsen P) and water extractable P (WXP).  Nitrogen mineralization rates were not significantly different from the control in three of the six treatments. The amended biochar (BA) at 10 Mg ha-1 increased N mineralization by 12 kg N ha-1.  Carbon mineralization rates in the BI decreased with increasing rates of biochar application. C-mineralized from BAamended were significantly greater than BI among all application rates. Significant increases in Olsen P were found at biochar application rates of 20 and 40 Mg ha-1. WXP was greater than the control soil at all biochar amendment rates.  Our research demonstrates that in Quincy sand biochar from anaerobic fiber coated with P from anaerobic digested effluent (ADE) may be a viable source of  alternative[MSOffice3]  P fertilizer.

Acronyms: ADF – animal digested fiber; ADE – anaerobic digested effluent; BI-Biochar; BA – 15 Day Coated biochar; WXP – water extractable P. 
Introduction

The demand for phosphorus (P), a macronutrient is projected to increase over the next 50 years in response to world population growth.  Analysts predict that the world supply of rock phosphate will be depleted within the same period (Cordell et al. 2009).  World demand for P is met by rock phosphate mined from three countries: USA, China, and Morocco.  As supply can no longer meet demand nations will potentially restrict rock phosphate exports.  As world supply and markets tighten, it will be critical to improve efficiency in plant uptake of P, and ways of recovering P from overloaded systems (Leikam and Achorn, 2005).   Phosphorus chemistry is complex and its cycling in the soil is a function of pH, temperature, soil structure, and clay mineral composition.  Phosphorus is either readily available to the plant via the soil solution, is bound to the soil matrix or can be mineralized from organic sources by soil organisms for uptake by plants (Bittman et al., 2004; Havlin et al., 2005).   Phosphorus is primarily made available to plants through mineralization of organic material and the breakdown of primary and secondary P minerals.  (Sharpley and Tunney, 2000; Pierzynski et al., 2005)

A problem with concentrated animal feeding operations (CAFO) and long-term dairy production is over-application of nutrients, specifically N and P (Sharpley et al. 2003).  The annual estimated dry manure produced in the United States by beef cattle, dairy cow and swine is 7.1 X 107 Mg which comprises 2.3 X 107 Mg of P. This amount is larger than the amount added annually by commercial fertilizer sources (Toor et al., 2006). Global demand for inorganic-P is expected to reach 38 Mt by 2014 (IFA). A dairy with 4000 Holsteins on site can potentially produce 1.1 X 107 L of liquid manure requiring disposal, removal and or utilization per year.  A common practice has been to apply waste lagoon water to adjacent agricultural fields (Hart et al., 1997), which in the long-term may cause environmental hazards from nutrient loading in the soil, runoff and eutrophication. Eutrophication is an increase in algae growth brought on by excess nutrient addition into an aquatic system, causing hypoxia and thus anaerobic zones as the micro and macro-biota consume dissolved O2 (Sharpley et al. 2003).  Low oxygen zones are problematic for fish and other aquatic life (Pierzynski et al. 2005; Sharply et al. 2000; Wendt and Corey, 1980).

One method for waste management and waste reduction in the dairy system is the use of anaerobic digesters (AD).   An AD system cycles solid or liquid waste through an environment optimal for methane production (Thauer, 1998). The benefits of an AD include reduced odor, reduced greenhouse gas emissions, and reduced solid waste (Kaparaju and Rintala, 2010; Hobson and Feilden, 1982; Rico et al., 2007; Gungor and Karthikeyan, 2008; MacConnell and Collins, 2009). However, the waste stream remains rich in nutrients, including P, making nutrient recovery necessary.  Knowledge of P recovery has advanced over the last decade in the wastewater industry (de Bashan and Bashan, 2004).  Methods have incorporated precipitation with metals (Thomas, 1972; Banu et al, 2008; Matsumiya et al, 2000; Bashan and Bashan, 2004; Delhaize and Ryan, 1995), microbiological conversions, ( Uusi-Kamppa et al., 2000) constructed wetlands, or a combination thereof (Nairn and Mitsch, 2000; Leader et al., 2005).   In response to the need for P removal in dairy systems, the formation of struvite (magnesium ammonium phosphate hexahydrate, NH4MgPO4-6H2O) has been revisited for P recovery (Westerman et al, 2009; Massey et al., 2007; Schuling and Andrade, 1999; von Munch and Barr, 2001; Bowers et al., 2007; Bowers and Westerman, 2005). However, the high concentration of Ca in dairy manure as it competes for sites with Mg complicates struvite production and therefore raises production costs. Struvite formation occurs within hours to weeks depending on the type of waste stream, P concentration, temperature, and concentrations of Mg and Ca. As an alterative to struvite production, researchers have investigated polymer flocculation; where small particles or suspended solids are flocculated and fall out of suspension (Frear et al., 2009, Ravina 1993). Chemical polymers can be added to the dairy effluent to induce flocculation for P removal, but the process is expensive. Furthermore, the optimization of the polymer flocculation process is difficult as the liquid chemistry is unique on any given dairy (Bashan and Bashan, 2004; Frear et al., 2009; Basu and MacDonald, 2007, Revina 1993).

There are few viable low-cost methods of P removal and recovery from dairy manure effluent.  Our research fills this gap by using biochar (charcoal) for P recovery.  Activated carbon (AC) from various sources has been used for P removal. Tamarind (Tamarindus indica Linn) nut shell AC removed 95% of the P from a standard solution through adsorption (Bhargava and Sheldarkar, 1993). Powdered AC from hardwood was used to remove P from industrial waste water effluent (Meidl, J.A, 1997).  Almond (Prunus amygdalus Batsch) and pecan (Carya illinoinensis) shells AC (Toles et al., 1997), and poultry-manure derived AC have been used to remove mercury from coal-fired power plants (Klasson et al., 2010; Yaji et al., 2008).  The use of AC is a proven method for containment and removal of P from waste streams. However, AC remains expensive, which has kept it out of dairy waste management programs.

The thermo-chemical conversion (pyrolysis) of plant biomass is being studied for its potential to produce second-generation liquid biofuels and the co-product biochar, which can be used as a soil amendment. Pyrolysis of biomass yields 60-70% of the original biomass as a bio-oil and 15-25% as biochar with the remaining 5-15% as ash (Graham et al., 1984; Czernik et al., 1994; Bridgwater et al., 1999; Mohan et al., 2006; Huber and Corma, 2007). Biochar is the carbon-rich product remaining after biomass has been heated at low temperatures (ca. 350°C to 600°C) in an environment with little or no oxygen (Amonette and Joseph, 2009). Biochar is currently the accepted term for pyrolysis-derived charcoal when used as a soil amendment (Sohi et al., 2010).

By using biochars to adsorb N and P in soils and from waste water streams, it is possible to add value to bio-char. Because biochars are significantly more stable than both the fast-and slow-cycling fractions of soil organic matter, biochar additions can have significant long-term benefits on soil C sequestration (Lehmann et al., 2006; Lehmann, 2007).  However, any value-added chemical processing treatment performed on the biochar to improve its nutrient adsorption capacity will increase production costs.

[MSOffice4] Beneficial characteristics of biochar as a soil amendment are its high cation exchange capacity (CEC; 40 to 80 meq per 100 g), high surface area (51 to 900 m2 g-1), which leads to increased water holding capacity, and affinity for micro- and macro- plant nutrients (Laird, 2008; Gaunt and Lehmann, 2008; Cheng et al., 2008; Novak et al., 2009; Roberts et al., 2010). Application of charcoal reportedly increases soil pH, and in acid soils decreases the concentration of Al, which often limits crop growth in the tropics (Hecht et al., 1982; Piccolo et al., 1997). When soil was amended with charcoal, annual crop yields increased by 100% or more (Steiner et al., 2007). In many tropical and subtropical soils, charcoal increases exchangeable bases, CEC, and nutrient availability (Piccolo et al., 1997; Glaser et al., 2002) decreases soil bulk density, and improves water holding capacity (Liang et al., 2006; Busscher et al., 2010; Novak et al., 2010). The addition of biochar to soil also increases soil C stocks that further improve nutrient storage and soils physical properties (Cheng et al., 2008).

The objective of this research was to assess the utilization of biochar for P recovery and determine whether the recovered P is plant available.  Authors found no current research addressing this subject.   We hypothesize that P amended biochar obtained from pretreatment with dairy effluent would 1) increase C-mineralization rates, 2) decrease N-mineralization due to the potential for immobilization, 3) increase inorganic P levels in the soil, and 4) release phosphate back into the soil profile making it available to plants.

Materials and Method

Feedstocks and Production of Biochar

Animal-digested fibers (ADF) produced after processing dairy manure through a GHD Plugged Flow (GHD Inc., Wisconsin) anaerobic digester (Outlook, WA) were collected as the biochar feedstock.  The ADF was air dried to less the 80g kg-1 moisture and commercially pelletized (Mid-Valley Milling Inc. Prosser, WA) at 207 MPa into 5 mm diameter pellets.

The biochar studied was produced by two methodologies the first was a batch pyrolysis reactor built by the Biological Systems Engineering Department at Washington State University (Granatstein et al., 2009). This unit could process 200 g of biomass. The batch reactor vessel was a 585-mm (Model HTF55322A) furnace. The reactor was heated to 500oC (heating rate: 25oC min-1) and kept at that temperature for 30 minutes. After each run the oven was turned off and the biochar was left inside the reactor to cool to ambient temperature. The pyrolysis vapors were evacuated from the reactor using 1 L min-1 (Normal Temperature and Pressure: NTP) of N2 as carrier gas.

Secondly, biochar was produced from the ADF pellets at a pyrolysis temperature of 500oC for four hours using a barrel retort batch unit design (Marv’s Reliable Maintenance, Bellingham, WA). Pyrolysis temperature of the retort was monitored with XCIB-K-1-6 thermocouples (Omega Incorporated, Stanford, CT) mounted inside the retort and connected to a Campbell Scientific X10 Data Logger programmed to continuously (2 minute) collect data for temperature verification (Chapter 3 Figure 1). Upon batch completion the retort was allowed to air cool until biochar was removed and stored dry.

Biochar and Soil Analyses

Bulk soil samples were collected from the top 30 cm of a native Quincy sand (sand, mixed, mesic Xeric Torripsamment) at the USDA-ARS Integrated Cropping Systems Research Field Station near Paterson, Benton County, WA (45o56’ N, 119o 29’ W; 114 m above sea level). Quincy sand is a young alluvial soil, with low nutrient and water holding capacity, occurring in central Washington and cropped in vegetable-based rotations under irrigation (Table 1).

The soil was amended with three rates (10, 20 and 40 Mg ha-1) plus a non-amended control of biochar (BI) and biochar (BA) amended with AD dairy effluent (Streubel Dissertation). Total soil organic C and N were determined by dry combustion on an elemental analyzer (CNS Vario EL III, Elementar, Hanua, Germany). A pH meter (Model 445, Corning Inc., Corning, NY) was used to measure the pH of each biochar and amended soil, a ratio of 10 g of soil or biochar and 20 ml of deionized water  was mixed together and the pH taken using the method of Robertson et al. (1999)  .  The mineral and organic matter fractions of each biochar (BI, BA) were determined by dry ashing 0.5 g of each char at 500oC for 24 h and recording the weight loss after ignition (Nelson and Sommers, 1996). The organic fraction was estimated by the difference between the initial sample weight and the ash remaining. Macro- and micro-elemental composition (P, K, S,  Ca, Mg and Cu),  of biochars was determined by inductively coupled plasma atomic emission spectroscopy (ICP; DV3300, PerkinElmer, Waltham, MA) according to Isaac and Johnson (1998).

Soil N mineralization potentials were determined over a 49-d incubation. Amended soil (10 g) was incubated at 70% of field capacity (FC) at 25oC and extracted weekly with 50 ml of 2 M KCl, shaken for 1 h on a rotary shaker, then filtered through a Type A/E glass fiber filter (Gelman Sciences Inc., Ann Arbor, MI). The extracts were analyzed colorimetrically for NO3‑N and NH4-N using a flow-injection analyzer (FIA) equipped with an auto sampler (QuikChem AE; Lachat Zellweger, Loveland, CO).

Water extractable (WXP) and Olsen P concentrations, were determined over a 21-day incubation.  Amended soils were incubated at 70% field capacity (FC) at 25o C.  Subsamples for WXP analyses were taken on days 0, 7, 14, 21, and days 0 and 21 for Olsen P. Briefly, water extractable P was collected by shaking 5 g of incubated biochar-soil in 50 ml of deionized water for 1 hour.  Immediately upon shaking samples were centrifuged for 30 minutes at 335 Hz.  Filtrate was then decanted and filtered using a Whatman 42 filter.  Sodium bicarbonate (Olsen P) was determined by the method of Kuo et al. (1996). The sodium bicarbonate P was determined by shaking 2 g of biochar-soil in 40 ml of 0.5 M NaHCO3 for 30 minutes then filtered using a Whatman 42 filter.  Extracts were stored at 5oC until analyzed using the ascorbic acid colorimetric method on an auto sampling spectrometer (EasyChem Analyzer, Systea Scientific, Oak Brook, IL).

Carbon mineralization was used as a measure of soil biological activity and stability of biochar C using the static incubation method (Zibilske, 1994). Soil samples amended with BI and BA (25 g) at each biochar application rate were adjusted to 70% FC [MSOffice5] in 160-mL containers and were incubated at 25°C for 30 weeks. The 70 % FC was consistent with the experimentation completed in the first part of this dissertation.  Headspace CO2 was measured weekly by direct injection of gas samples into an infrared gas analyzer (IRGA; type 225- MK3, Analytical Development Co., Hertfordshire, UK). After each analysis, samples were adjusted for moisture loss and the headspace was returned to ambient by degassing with compressed air.

Statistics

   Data were tested for normality by Shapiro-Wilk’s W test and for heterogeneity of variance by Levene’s test using SAS (version 9.2, SAS Institute, Cary, NC). Associations between biochar and 15 day coated biochar, WXP P, Olsen P, and C and N mineralization were assessed using PROC GLM and ANOVA with pairwise comparisons of LSMEANS by Tukey at P < 0.05 (version 9.2, SAS Institute, Cary, NC).

RESULTS AND DISCUSSION

Elemental Concentrations of Biochars

The BI and BA had C concentrations of 483 and 462 g kg-1, respectively (Table 2). The N concentrations were similar in both biochars at 22 g kg-1. The sulfur content average 56 g kg-1 for both biochars used in the experiments. The characteristics of the biochar are largely dependant on the original feedstock, temperature and pyrolysis system.  The biochar used in this experiment was characterized by Streubel et al., (2011;Dissertation).  They found the BA had higher concentrations of inorganic P than BI .   Streubel et al. (2011; Dissertation) found the loss of C and N during pyrolysis was related to the original feedstock concentrations of ash, lignin, cellulose, and hemicellulose and the temperature of pyrolysis.  Further characterization of the chemical composition of each biochar can be found in this dissertation.

The ash and organic content were slightly different between the BI and BA biochars with a lower mineral and higher organic fraction (Table 2). The organic fraction was greater in the BA than BI reflecting the adsorption of soluble C and fiber present in the dairy effluent. Ash contents were greater than has been reported for woody feedstocks but were similar  to herbaceous feedstocks and manures  (Shinogi and Kanri , 2003).  Shinogi and Kanri (2003),  concluded that C, N, H, O, and S are lost during heating as the inorganic salts that comprise the ash are not volatilized.  In a study of ash elemental composition from six different wood sources their total P concentrations ranged from a high in Red Oak of 15.6 g kg-1to a low of 5.6 g kg-1 in Western Oak (Misra et al., 1993).  The P concentrations were higher in the BA (10.7 g kg-1) then BI (9.2 g kg-1). Garg and Bahl, (2008) found cow manure to have a P content of 8.5 g kg-1.   The characterization of the forms of P on the two biochars was described (dissertation).  The K concentration was similar between the BI (15.4 g kg-1) and BA (14.8 g kg-1) The BA was immersed in effluent for 15 days where leaching of K could have occurred. The concentration of Cu was also higher (40%) in the BA than BI this is most likely resulting from mineral supplements in the dairy cattle feeds and the Cu foot baths used to treat hoof infections. Cao and Harris (2010) found similar elemental concentrations and trends in dairy manure.

luence of Biochar Additions on Soil Properties

Soil N Mineralization

There was a consistent slight increase over the control in the N mineralized with the addition of biochar regardless of application rate but it was not significant (Figure 1). The N mineralized decreased with the BA additions at the two highest application rates but showed a significant increase at the 10 Mg ha-1 rate (Figure 1). The lower N mineralization rates for the BA may be due to the higher concentration of organic matter (OM) present on the biochar resulting in immobilization of N as the soil microbial populations decompose the OM. Other studies have shown that lower N mineralization may be related to the biochars’ concentration of volatile material that inhibits microorganisms in the soil and thus decreases NO3 production (Deenik et al., 2010). Switchgrass residue added with biochar to a sandy soil could potentially stimulate the availability of mineralized N (Novak et al., 2009). This could account for the significant difference in the BA at 10 Mg ha-1 rate where concentrations of C were not high enough to promote N immobilization.  Biochar amendment has also been shown to decrease N availability in some Amazonian soils (Lehmann et al., 2003), which has been viewed as beneficial sincef  fixation of N by biochar would reduce losses by leaching. Others report as much as 70% less N leached from soils that had been amended with biochar (Laird, 2008; Cheng et al., 2008). Our data was consistent with the range of results published in the literature and found in our early studies (Streubel et al., 2011; dissertation)

C Mineralization

The C-mineralization coefficient, the percentage of total organic C evolved as CO2, has been used to determine the effect of soil and cropping systems management on the decomposition and cycling of SOM (Collins et al., 1992, 1999; Paul et al., 1999, 2006). Long-term incubations (>200 d) have also been used to differentiate functional C pools in soil (Motavalli et al., 1994; Paul et al., 1999; Cochran et al., 2007; Collins et al., 2010).

The rates of C mineralization measured among biochars in the 217-d incubations are greater for BA than BI (Figure 2).  These data are consistent with values reported in other studies (Bruun et al., 2008; Zimmerman, 2010; Novak et al., 2010). The C‑mineralization rate curves were similar for the two treatments and three rates (Figure 2), however the BA was higher than the BI and Quincy sand control.  The BA 40 Mg ha-1 rate was significantly higher then all other applications. C-mineralization was significantly greater in the initial days of the incubation which can be partly explained by the presence of labile C on the biochars (Smith et al., 2010). The higher rate of mineralization found in the BA is explained by the AD effluent coatings on the biochar which would be easily accessed by microbes in the soil.  The coatings on the BA ranged from 5-8% (data not shown). The increase in CO2 evolved from The BA ranged from 6 to 15 mg CO2-C kg-1 soil d-1, depending on the treatment and rate. Under low pyrolysis temperatures (250 to 600oC), the thermo-chemical reaction between pyrolysis and liquefaction can be variable (Demirbas, 2000).  At 500oC, Keiluweit et al. (2010) found that the volatile matter content of biochar was still fairly high in fescue C but had dropped significantly for a Pinus ponderosa biochar. Smith et al. (2010) showed that newly produced biochar added to soils increased C-mineralization for the first six days of incubation, and then rates returned to pre-addition levels. Using δ13C analysis they also showed that the increase in CO2 efflux originated from biochar-C and not from the native SOM. Our results with BA treatments remained higher than the controls over the 217 d incubation but did decrease over time.

The rate of biochar addition had only minor effects on total C mineralized, however presence of coatings on the biochar was significant (Table 3). The total C mineralized decreased as the rate of biochar addition increased, similar to that observed by Bruun et al. (2008) and Zimmerman (2010). The reduction in C-mineralization results from dilution of soil organic C with C that is largely biologically inert. The opposite was true for the BA additions which increased with rate.  In contrast to the biochar the increase of C-mineralization is the addition of biologically available C from the coatings in the biochar.  Our previous results (dissertation) have shown the recalcitrance of biochar, however the addition of coatings from dairy effluent significantly increased the total cumulative C-mineralization.

Phosphorus in the Soil         

Olsen P increased over the soil control within the highest and lowest application rates of BI and BA biochars (Table 4).  The Olsen P extraction is considered useful for giving growers an index of available P for the entire growing season, therefore little measureable change is expected due to the short duration of the  laboratory incubation. The 20 Mg ha-1 rate is similar throughout all treatments.  Halanjnia et al. (2009) found that Olsen P decreased over 150 days in soils that were amended with manure.  They suggested that the elements such as Fe, Ca, Mg, and Mn could be binding P to the soil complex.  This explanation could be plausible given the biochar is made from and is coated with dairy manure; however why the decrease is only consistent at the middle rate needs to be researched further.  Garg and Bahl, (2008) found when adding farmyard manure to a Samana sandy loam the Olsen P increased 8 mg kg-1.  Similarly in biochar Chan et al. (2007) also found an significant increase in extractable P from an Alfisol amended  with greenwaste biochar at rates exceeding 50 Mg ha-1  in a greenhouse pot trial. . In a follow up experiment using poultry litter derived-biochar produced at 450 and 500 Co the extractable P concentrations at 10 – 50 Mg ha-1 also increased. The increase was attributed to the C – P complexes in the biochar that formed under the different temperatures (Chen et al., 2008). The significant increase in the Olsen P in the crushed biochar could be explained by a more uniform distribution of biochar in the soil then by the pellet.  The lack of a consistent result as application rates increased maybe due to the initially high soil content of Olsen P. This condition diluted out the P incorporated by the BA biochar since at the high rates of amendment only 1.6 mg P was added in a background of 15 mg P extractable P.   Our results were both supportive and contrary to the literature. However, this study is the first to use  these types of biochars. Further investigation is necessary to clarify the observed trends in P availability from dairy effluent amended biochars.

Water extractable P showed an immediate increase in the biochar amended soil over the un-amended soil in all but two rates and treatments, however not all increases were significant (Table 5).  The BA at the highest rate was significantly higher than all treatments for the 21 day period.  The BA at the highest rate increased the average WXP from 4.17 to 12.4 mg kg-1 over the control (Figure 4).  The pattern of increase was BA-crushed 40 Mg ha-1 > BA crushed 20 Mg ha-1 > BI-crushed 40 Mg ha-1 > BI-crushed 20 Mg ha-1 > BA crushed 10 Mg ha-1.  The pelleted BI and BA chars did not significantly increase the WXP of the control soil. This is most likely because the pellet form has limited access to the entire sample volume where as the crushed biochars were uniformly mixed throughout the soil.

The WXP is a measure of what is immediately plant available in the inorganic PO4 pool.  Our data is similar to studies that have shown that in dried manure WXP represents the largest concentration of P (He et al., 2009; He et al. 2007; McDowell et al. 2007).  In this dissertation it was determined that there is a significant amount of inorganic P coated on the biochar surface. Solla-Gullon et al. (2006) found an increase of available P when wood ash was amended at 10 and 20 Mg ha-1.  Schiemenz et al. (2010) found an increase of from 8.6 to 12.4 mg kg-1 of WXP when straw ash was amended to the soil amended with 9.8 g per 6 kg. They found there was not a difference between commercial P and the straw ash when it came to WXP in the soil. Charred bark from the Acacia mangium tree increased WXP from 4.5 to 5.4 g kg-1 in the soil (Yamato et al., 2006).  The increase from cow manure alone in the WXP fraction was 3.2 mg kg-1 after six years compared to the control soil.  They concluded there were both inorganic and organic forms that were rapidly mineralized in the soil or transformed into longer term liable forms (Eichler-Lobermann et al., 2007).

Cao and Harris (2010) had P concentrations of 26.6 g kg-1 in 500oC manure biochar and believed it was because of the crystallization of amorphous P-Ca-Mg which is less soluble in the soil solution compared to the water extractable form of P.  This was confirmed by the presence of whilockite ((Ca,Mg)3(PO4)2) in their manure biochars at 500oC. Gundale et al. (2006) confirm ed that Ca and Mg will most likely remain in the biochar at 500oC while elemental P will begin to volatilize at that temperature. This serves as a likely explanation to our results seen in both the Olsen P (Figure 3) where B had the highest extraction of P but in the WXP fraction the BC15 1.5 Mg ha-1 had the highest extraction of P (Figure 4).  The coatings serve as the WXP fraction and the Olsen P fraction was incorporated into  the biochar structure in the form of P-Ca-Mg.

Conclusion

The concerns over excess P in the environment and its ability to cause eutrophication, especially around CAPO’s, has made new P recovery methods necessary.  One such method is the use of biochar as a filtration media for P recovery in anaerobic digested effluent. Charcoal has been used as a soil amendment in humid equatorial climates for centuries. Biochar amendments have the potential to improve temperate soils by adding C, raising pH, and increasing WHC. Biochar after being cycled in AD effluent is an alternative end-product for local farmers needed to reduce phosphorus loading. The amendment of Quincy sand with biochar pyrolyzed from anaerobic digested fiber and then used to filter AD effluent had mixed effects on soil N and C mineralization and pools of extractable P. The rates of N mineralization showed an increase with 10 Mg ha-1 in the BA and in the remaining treatments N mineralization decreased. With the addition of the untreated biochar, the percentage of total C mineralized decreased as the rate of biochar amendment increased, however the addition of the BA increased C-mineralization due to the presence of labile C on and within the biochar matrix.   Soil Olsen P levels increase over 21 d in the biochar amended soils at the higher rates but tended to decrease in the BA amended soils.  In contrast, the WXP fraction showed an increase at all rates of BA additions and only the highest rate of BI.  The levels remained elevated throughout the 21 d period.  The use of P coated biochar as a soil amendment for a P source appears to have potential.  The recalcitrant nature of biochars can also improve C sequestration in agricultural soils but further studies should be conducted to determine effect of the proportion of total C in biochar in the soils when the coating is added.  We found that biochar coated with P from effluent is a significant factor in raising WXP in Quincy sand. However, further long term research is needed to better understand the effects.  Field-scale and lab research is also needed to understand interactions between the coated biochar P, the soil matrix and plant interactions.

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Table 1. Selected characteristics of a Quincy sand used in the laboratory incubations.

Soil

Soil Characteristics

Series Texture

C

N

S

C:N C:S pH CEC
 ————— g kg-1 —————— cmol kg-1
Quincy Sand   4.3 (0.5) 0.5 (0.1) 0.2 (0.03) 8.6 22 7.1   3.3

Add footnote describing symbols and an explanation that the parentheses contain standard deviations.

Table 2. Elemental composition of the untreated and dairy effluent amended biochars.

C

N

S

Ash

Organic Content

—————————————g kg-1————————————–

BI

483.5 (1.2)

21.7 (0.07)

5.5 (0.03)

431.6 (1.2)

568.4 (1.2)

 BA

461.9 (0.3)

22.0 (0.07)

5.7 (0.03)

422.5 (0.2)

577.5 (0.2)

P

K

Ca

Mg

Cu

—————————————g kg-1————————————–

BI

  9.2 (0.3)

15.4 (0.4)

28.8 (2)

11.1 (1.7)

 0.06 (0.01)

 BA

 10.7 ( 0.8)

14.8 (1.3)

   38.2 (0.8)

12.2 (0.3)

0.10 (0.3)

BI- un-amended  biochar.  BA- amended with dairy effluent.  Std. deviation in parentheses.

Add footnote with abbreviations for the elements.

Table 3.  Total cumulative C-mineralization for a Quincy sand soil with

untreated and dairy effluent amended biochars.

C    Mineralized

Soil/

BiocharChar

RateCumulative

CO2-C

C-min difference

From control

 g kg-1

—————– mg kg-1 —————–

Quincy0

518.67 (43)d

0

 BI4.0

502.07 (30)d

-16.6

 7.5

519.72 (19)d

1.05

 15.0

457.96 (23)cd

-60.71

 BA4.0

542.04 (43)abc

23.37

 7.5

593.54 (42)ab

74.87

 15.0

662.37 (34) a

143.7[MSOffice6]

BI- un-amended biochar.  BA- amended with dairy effluent.  Values in

parentheses are standard deviations. Values followed by the same letter

are not significantly different at P=0.05 n=4.

Table 4. Statistical analysis of Olsen P for 21 day soil

incubations with untreated and dairy effluent amended

biochars.

TREATMENT

RATE

Mg ha-1

mg kg-1 P

LSMEAN

BI- crush

40

9.59

A

BI- crush

10

9.44

B

BA-pellet

40

8.55

ABC

BA-pellet

10

8.28

ABC

BA-crushed

40

7.67

C

Quincy soil

0

7.49

C

BA-crushed

10

7.35

C

BI-pellet

10

7.30

C

BI- crush

201

6.97

C

BA-crushed

20

6.95

C

BI-pellet

40

6.81

C

BI-pellet

20

6.45

C

 BA-pellet

20

6.26

C

Different letter represent significant differences

at p=0.05 for Tukey’s LSmeans.

Table 5. Statistical analysis of water extractable P

for 21 day soil incubations with untreated and dairy

effluent amended biochars.

Treatment

RATE

Mg ha-1

LSMEANS

BA-crushed

40

12.47 A

BA-crushed

20

8.02 B

BI-crushed

40

7.82 BC

BI-crushed

20

7.11 BC

BA-crushed

10

6.81 BC

BI-crushed

10

5.91 C

BA-pellet

20

5.64 C

BA-pellet

40

5.13 C

BI-pellet

40

4.37 C

Quincy soil

0

4.17 C

BI-pellet

10

3.97 C

BI-pellet

20

3.83 C

Different letter represent significant differences

at p=0.05for Tukey’s LSmeans

Figure 1 Total NO3-N Mineralization after Day 49 in Quincy Sand

QS – Quincy sand

B – ADF Biochar

BC – 15 day coated biochar

Different letters represent 0.05 significant differences

Figure 2 Total rate of CO2-C for APF Biochar and APF Biochar with Coatings at Three Rates of Application in a Quincy Sand Over 217 Days Letters represent

significance at >0.05

Figure 3. Sodium Bicarbonate extractable P from three amendment rates of untreated and dairy amended biochars.

Figure 4 Water extractable P from three amendment rates of untreated and dairy amended biochars.

CHAPTER FIVE

GENERAL CONCLUSIONS

Biochar is an alternative end-product for local farming and forestry residues such as dairy waste, forage residues, forestry slash, and other production debris. Amendment of temperate soils with biochar has the potential to improve soil health by increasing  soil C,  the pH, and WHC. Amendment of agricultural soils with biochar pyrolyzed from herbaceous and woody feedstocks had variable effects on soil properties depending on soil type, biochar feedstock, and rate of amendment. In the Hale silt loam and Quincy sand, rates of N mineralization increased among some amendment rates and biochar sources, whereas in the remaining silt loams N mineralization tended to decrease. In the Quincy sand, Palouse, Thatuna, Naff and Hale silt loams amended with biochar, the percentage of total C mineralized decreased as the rate of biochar amendment increased, and there was a linear relationship between the rate of biochar amendment and the change in C content of the soil. Soil pH increased with biochar amendments among all soil types and biochar feedstocks. The use of biochar to raise soil pH may be beneficial where long-term fertilizer applications have acidified soils. Conversely, carbonate-rich soils with high pH that are common in arid regions may not benefit from additions of biochar because increases in soil pH can be detrimental to micro-nutrient availability and to crop production. Soil water holding capacity varied among rates of amendment and biochar feedstock. In irrigated systems, appropriate choice of biochar feedstock could increase soil WHC and thus reduce either the frequency or amount of irrigation. The recalcitrant nature of biochars can also improve C sequestration in agricultural soils because the proportion of total C in biochar that is recalcitrant leads to long MRT in soil. We found that biochar feedstock is not a significant factor in raising pH or C in five regional soils.

The use of biochar by the dairy industry may produce significant environmental benefits.

We recovered 30 % of the P from dairy effluent over a 15 d period using biochar from anaerobically digested dairy waste fiber.  The P recovery is dominated by the water-extractable forms that dry onto the biochar.  Water and Olsen P extractions showed significant increases of plant-available P on the biochar.  The liquid and solid state 31-P NMR confirmed the presence of the inorganic species of P on the biochar, whereas Ca-P complexes dominated the solid phase.  The NMR did not quantify P concentration. While biochar has been studied for its soil impacts, this research is the first data representing biochar produced from ADF and used to recover P from dairy effluent.  Although recovery was less than anticipated, we demonstrated a sound concept; a 32 % recovery of P from effluent represents a benefit to both the environment and dairies.

Once amended to soil, the P-amended biochar (BA) increased N-mineralization at low amendment rates and tended to decrease N-mineralization at higher rates.  In non-amended biochar, the percentage of total C mineralized decreased as the rate of biochar amendment increased. Soil Olsen P levels tended to increase over 21 d with biochar amended soils at higher rates but tended to decrease in the BA amended soils.  By contrast, the WXP fraction increased at all rates of BC15 additions and only at the highest rate of non-amended biochar. The rates remained high throughout the 21 d incubation.  We found that biochar coated with P from effluent was a significant factor in raising WXP in the Quincy sand.

This research indicates that biochar recovery of P from dairy effluent is a viable concept. However, further long-term research is needed to better understand the effects of the recovered P on the biochar itself and in the soil once amended.  Field-scale and laboratory research is needed to understand the interactions among the P amended biochar, the soil matrix, and crop plants.


 [MSOffice1]Not in final format. I assume because you will be making corrections.

 [MSOffice2]rewrite

 [MSOffice3]Available?

 [MSOffice4]Much of this introduction is in the materials and methods but does not read as clearly as the version in this chapter. I would replace the similar comments in the introduction with these

 [MSOffice5]Add a sentence explaining the choice of the this FC and not another.

 [MSOffice6]Adjust to have the same significant figures

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