Chemical Engineering Journal 279 (2015) 3137

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Chemical Engineering Journal

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Ammonia recovery from anaerobic digester effluent through direct aeration

Quan-Bao Zhao a, Jingwei Ma a,, Iftikhar Zeb a, Liang Yu a, Shulin Chen a, Yu-Ming Zheng b, Craig Frear a

a Department of Biological Systems Engineering, Washington State University, Pullman, WA 99164, USAb Institute of Urban Environment, Chinese Academy of Sciences, 1799 Jimei Road, Xiamen 361021, PR China

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

Solely injecting air into digester effluent can elevate pH and remove ammonia.

Smaller bubble size favors the process. Carbonate and ammonium concentrations first increase then decrease during aeration.

a r t i c l e i n f o

Article history:

Received 13 October 2014

Received in revised form 21 April 2015 Accepted 23 April 2015

Available online 29 April 2015

Keywords:

Aeration

Ammonia

Anaerobic digestion

Dairy manure

Modeling

a b s t r a c t

A direct aeration strategy was developed for ammonia recovery without alkali addition. The main chem-ical buffer system of digested effluent is made up of high concentration of acidic species, which lead to a hypothesis that elevation of pH for ammonia release could be achieved by removing supersaturated CO2, dissolved carbonate, and bicarbonate solely. The concept was tested in lab scale reactors with digested dairy manure focusing on temperature, bubble size, liquid depth and airflow rate. It has demonstrated that simple air stripping without alkali chemical input is an effective way to elevate pH of the digested effluent due to intriguing chemical shifts strongly related to the high levels of carbon dioxide, bicarbon-ates and carbonates present in digested effluent. These chemical shifts, ultimately release carbon dioxide and raise the pH of the effluent to levels near 10, which with combined elevated operating temperatures from waste engine heat, can lead to 7090% shift from ionic to free, gaseous form of ammonia and sub-sequent recovery of ammonia through acid contact. The chemical relationships and equilibrium shifts associated with the aeration process and its subsequent release of gases were further investigated by chemical equilibrium model.

2015 Elsevier B.V. All rights reserved.

1. Introduction

Farm-based anaerobic digestion (AD) systems offer multiple benefits such as renewable energy, methane destruction, carbon avoidance, and value-added production of digested fibrous solids [1]. These systems typically produce a nutrient-rich effluent that

Corresponding author. Tel.: +1 509 335 0194; fax: +1 509 335 2722. E-mail address: mjw@wsu.edu (J. Ma).

http://dx.doi.org/10.1016/j.cej.2015.04.113 1385-8947/ 2015 Elsevier B.V. All rights reserved.

still requires proper disposal to land, with large animal operations producing voluminous amounts of wastewater that hold potential for overloading nearby soils with excess nitrogen and phosphorous [2]. Technological responses to these nutrient concerns has simultaneously grown in interest, with particular attention tuned to the development of nitrogen technologies that can either, alone or in series with AD, reduce nitrogen content in the wastewater [3]. An AD research focus has been on ammonia recovery and/or removal as the AD process converts a portion of organic nitrogen32Q.-B. Zhao et al. / Chemical Engineering Journal 279 (2015) 3137

to inorganic ammonia, thereby increasing total ammonia nitrogen (TAN) concentration in the wastewater.

Ammonia removal can be accomplished biologically by converting ammonium into non-reactive nitrogen gas through nitrification/denitrification [47] or anaerobic ammonium oxidation (Anammox) [8]. Ammonia recovery processes, which allow for pro-duction of potential agricultural fertilizers, include ammonia stripping-absorption [913], struvite precipitation [14,15], and membrane (gas permeable and reverse osmosis (RO)) separation [16,17].

Biological nitrification and denitrification processes are unlikely as a costeffective choice, as there are not enough bio-available organic materials in digested manure. The process also requires large reactors and energy inputs [47]. From a sustainability and economics perspective, the nitrogen is lost in the process to the air in nonreactive form, producing no saleable co-products [3]. Crystalline struvite in the form of magnesium ammonium phos-phate is an attractive slowrelease fertilizer but with limited nitrogen content. However, because of the low amounts of magnesium relative to ammonia concentrations, inputs of magnesium reagent are often required for the treatment of digester effluent, which can result in an expensive process [18]. Moreover, the majority of phosphorus in anaerobically digested effluent is tied up in a fine suspended calciumphosphate solid, thus becoming unavailable for struvite formation [19]. Membrane processes, while useful in obtaining greater levels of wastewater purification, require costly chemical pH treatment, either through prior acidification when utilizing RO or alkaline insertion for fast transportation of free ammonia through gaspermeable membranes [16,17]. In addition, membranes might require extensive pretreatment of the manure wastewaters, which are typically high in suspended solids content, while also being prone to high energy and operational and mainte-nance inputs [20].

Ammonia stripping is a potentially applicable process as there is a considerable concentration of ammonia in the effluent due to the biological conversion process of AD. A saleable ammonium salt fertilizer product can be produced from subsequent absorption of the stripped ammonia with an acid [9]. Unfortunately, from a cost perspective, traditional ammonia air stripping requires either addi-tion of alkali chemical to elevate the pH or heat to increase the temperature to release the free ammonia [9,21,22]. Additionally, traditional stripping tower systems utilizing packing media or tray towers are prone to solids interception as the dairy AD effluent often contain high contents of solids [2].

A challenge that needs to be overcome in order to raise the pH is the high buffer capacity of digested effluent due to the presence of significant bicarbonate/carbonate, phosphate, and ammonia buf-fers, with total alkalinity (TA) of digested effluent being elevated by 5060% [2]. Dissolved inorganic carbon (DIC), TAN, TA and pH are interrelated to the concentration and transformation of supersaturated CO2 produced during AD with dissolved carbonate, bicarbonate, ionized ammonium and free ammonia mutually related within complex equilibriums. Appreciable levels of calcium, magnesium and phosphorus in effluent from dairy manure digestion also contribute to a high alkalinity [15,19] and these intricate relationships. Detection of concentrations of each species is difficult, requiring chemical equilibrium modeling from a few

Table 1

Properties of the raw dairy manure and digested dairy manure. simple input species for prediction of the rest of the other chemical species.

The hypothesis for this paper is that a sufficient pH could be achieved allowing for ammonia to be efficiently recovered without alkali chemical input, requiring only airstripping, elevated temperature supplied by recovered waste heat, and without use of a complicated stripping tower. A laboratory study was conducted to verify the feasibility of this strategy by investigating several limiting factors including temperature, bubble size, liquid depth and air flow rate. A chemical species modeling effort was completed and the chemistry behind aeration was elaborated through model for optimized recovery of ammonia from dairy AD effluent.

2. Materials and methods

2.1. Anaerobic digested manure

Anaerobic digested manure was collected from a commercial dairy digester in Lynden, WA. The digester was operated under mesophilic condition (38 LC) with a hydraulic retention time (HRT) of 22 days. The digester effluent was screened through a 0.5 mm screen to remove coarse fiber before the experiments. The detailed physical/chemical characteristics of the undigested and digested manure are listed in Table 1.

2.2. Experimental setup and operation

2.2.1. Effect of temperature on ammonia stripping

Experiments were conducted in a 500 ml Pyrex bottle with a working volume of 400 ml. The aeration bottle was placed in a water bath with autocontrolled temperature (ISOTEMP 1013D, Fisher Scientific, Pittsburg, PA). The temperature was set to 35, 55, and 70 LC, respectively for different tests. Air was pumped into the bottom of the bottle, through an air stone (Aqua-Mist, Hauppauge, NY) by a peristaltic pump (Masterflex L/S 7524-40, Fisher Scientific, Pittsburg, PA) at a constant flow rate of 400 ml/min. Before the aeration started, AD effluent was preheated for one hour to obtain the desired temperature. Outlet gas samples were taken with a 35 ml syringe and stored in 12 ml vacuumed borosilicate vials (Exetainer, Labco Limited, Wycombe, England). Ten ml of liquid sample was taken each time with a 5 ml pipette.

2.2.2. Effect of bubble size, liquid depth and air flow rate on ammonia stripping

Experiments were conducted in two 5-gallon column reactors containing either a coarse bubble diffuser (EPDM flex cap diffuser, Mooers Products Inc., Milwaukee, WI) or a microbubble diffuser (SSI-AFD270 (900)-EPDM Membrane Disc, Stamford Scientific International, Inc., Poughkeepsie, NY) installed at the bottom of the reactor. The reactor chamber was temperature controlled at 40 2 LC. Air was pumped to the reactor through a blower (Sweetwater S-41, Aquatic Eco-Systems Inc., Apopka, FL) and the flow rate was controlled by a flow meter (GF-64511250, Gilmont Instruments, Barrington, IL). Three liquid depths (2, 5, and 10 cm) and four air flow rates (5, 10, 20, and 35 LPM (l/min)) were tested with sampling procedures following those described in Section 2.2.1.

TAN (mg/L)DIC (mg/L)Alkalinity (mgCaCO3/L)pHDissolved CO2 (mL/L) aDairy manure1.760 95984 278.960 4606.95 0.14527 104Digested dairy manure1.443 27845 1014.230 8537.80 0.19846 121

a Tested from the collected gas by applying vacuum of 27 in Hg for 2 h at 55oC.Q.-B. Zhao et al. / Chemical Engineering Journal 279 (2015) 313733

2.3. Chemical analytical methods

The pH, alkalinity, and COD were analyzed according to Standard Methods [23]. Content of CO2 in gas were determined by a gas chromatograph (Varian, CP-3800) [24]. Total ammonia nitrogen (TAN) was analyzed according to Standard Methods [23] using a Tecator 2300 Kjeltec Analyzer (Eden Prairie, MN, USA). The DIC was analyzed with a TOC-500 analyzer (TOC-500, Shimadzu, Kyoto, Japan) after centrifuge at 10g for 5 min and fil-tered through a 0.22 lm membrane. Total Phosphorus (TP) was analyzed using the Hach PhosVer 3 ascorbic acid method with acid persulfate digestion.

2.4. Chemical equilibrium model

The chemical equilibrium shifts of carbonate/bicarbonate and ammonia/ammonium were simulated by model, which considered pH, temperature, salinity, and concentration of dissolved inorganic carbon and TAN as described in Eqs. (1)(8) as listed in Table 2. Reaction (4), (5) and (7) are instantaneous while reaction (3) is slow [25]. The equilibrium constants of the carbonate/bicarbonate system were referred from a seawater CO2 model [26] and the concentration of each carbon species was calculated with MATLAB code as reported by Zeebe and WolfGladrow [27]. The pKa of ammonia/ammonium was calculated based on equation derived by Emerson, Russo, Lund and Thurston [28], then the concentration of [NH3] and [NH4] were calculated following Eqs. (1) and (2):

NH3& NH3 NH4&NH3 NH4&9

1 H&=Ka

1 10pKa _pH

NH4& NH3 NH4& _ NH3&10

where [NH3] is the free ammonia concentration (mol/L), [NH4] is the ionic ammonium concentration (mol/L), [NH3 NH4] is the total ammonia concentration (mol/L), [H+] the hydrogen ion concentration (mol/L), and Ka is the acid ionization constant of ammonia (mol/L). pKa can be expressed as function of temperature T (K) by the following equation [28]:

pKa 0:09018 2729:92=T11

The experimental data from Section 2.2.1 at temperature of 55 LC was used for model simulation. DIC, pH, and salinity served as model inputs and concentration of CO23_, HCO_3, CO2 (aq.) and CO2 partial pressure (pCO2) were calculated as model output. The error between model simulation and experimental data was eval-uated with pCO2 ( Fig. 3D).

Table 2

Equilibrium equations associated with ammonia stripping.

No.EquationpKa aRefs.(1)CO2 aq $ CO2 g

(2)H2CO3 $ CO2 aq H2 O

(3)HCO3_ H $ H2 CO35.82 [26](4)CO32_ H $ HCO3_11.17 [26](5)H OH_ $ H2 O6.58 [35](6)NH3 aq CO2 $ NH2 COO_ H

(7)NH4 $ NH3aq H8.41 [28](8)NH3 aq $ NH3 g

a pKa was calculated at temperature of 55 LC. 3. Results and discussion

3.1. Effect of temperature on ammonia stripping

The pH required for ammonia stripping decreases with increasing temperature ( Fig. 1A). For example, the pH required for shifting 70% of TAN to free ammonia at temperatures of 35, 55, and 70 LC are 9.32, 8.78, and 8.41, respectively ( Fig. 2B). The chemical shift between NH3 and NH4 is dependent on a number of factors in addition to initial TAN concentration. Most important among these are pH and temperature, as the concentration of NH3 increases with increasing pH and with increasing temperature [28]. The pKa of the ammonia/ammonium system is 8.95, 8.41, and 7.93 at temperature of 35, 55, and 70 LC, respectively. Increasing temperature enhances the molecular diffusion coefficient of ammonia in both liquid and gas film, and also will reduce the liquid phase viscosity and surface tension of liquid phase and the liquidgas distribution ratio of ammonia [28]. Increasing temperature has a very important effect on desorption of ammonia from water because of increasing dissolution. The increase of temperature will also cause increase in the rate of mass transfer and KLa [29]. The high correlation between temperature and the amount of ammonia removed was also observed by Liao, Chen and Lo [22]. They concluded that higher temperatures in the batch stripping of swine manure increased the removal efficiency. It was reported that NH3 volatilization rate could be elevated by 5-fold when the wastewater was heated to 60 LC [30]. Walker et al. showed that temperature had more impact on ammonia removal than air flow rate and aeration time [31].

From this study it can be noted that under constant aeration rate with no alkali addition, pH elevations necessary for TAN removal were achieved at the respective temperatures, with rate of pH elevation, maximum pH achieved, and ultimate TAN removal all higher with elevated temperature ( Fig. 2C). It took about 3, 4 and 12 h for 70, 55 and 35 LC, respectively, to reach their respective highest pH values under the same airflow rate. From a TAN removal perspective, 90%, 40% and 20% removal of TAN were achieved at the end aeration for 70, 55 and 35 LC, respectively ( Fig. 1D). TAN removal slowed down after the first few hours, which might due to a decreasing driving force as according to Henrys law the reduction in free ammonia concentration over time would slow the dissociation of ammonium salts (e.g. struvite).

3.2. Effect of liquid depth, bubble size, and air flow rate on ammonia stripping

3.2.1. Liquid depth

Contradictory to conventional wisdom that increased air travel time in liquid can assist in ammonia removal, this study found that reduced liquid depth, i.e. short bubble travel time, enhanced pH elevation and ammonia desorption ( Fig. 2A and B). The rate of pH elevation and achievement of final pH was higher for low depth aeration. The final pH for the low depth (5 cm, equivalent to of whole depth) reactor reached as high as 9.29.4 within 4 h, while pH in the high depth (10 cm) reactor only raised to around 9.0 after 6 h ( Fig. 2A). Liquid depth also had notable effect on ammonia removal rate. After 12 h aeration, the TAN removal rate in low and high depth reactors reached 80% and 5060%, respectively ( Fig. 2B). In the tested range of this study, air to liquid ratio is not an independent parameter, and the ammonia stripping process was more controlled by liquid depth than airflow rate.

Driving force is expressed as the difference between the bulk and interface concentration, with the former greater in standing than the latter for a positive driving force. In a column reactor, a positive CO2 driving force, due to low CO2 partial pressure in air,34Q.-B. Zhao et al. / Chemical Engineering Journal 279 (2015) 3137

Fig. 1. Effect of temperature on changes of pH and TAN removal from the AD effluent.

promotes fast CO2 release to air bubbles, leading to a rapid pH ele-vation for ammonia equilibrium shift in favor of ammonia desorp-tion. As bubbles travel along the column from bottom to top, the driving force for both CO2 and ammonia decrease as their concen-tration increase in the air bubbles. At some point, the driving force becomes zero or negative, making high liquid depth useless or even unfavorable for ammonia stripping. Reduced liquid depth takes advantage of a positive driving force and makes the stripping process more efficient. On the other hand, high depth requires increased stripping pressure, resulting in a compromised driving force and extra operating cost for the system. However, low liquid depth could lead to a large surface area requirement for the reac-tor, which could be compensated by reducing liquid retention time by using fine bubble size or increasing air flow rate.

3.2.2. Bubble size

Curves for pH elevation and ammonia removal were differenti-ated into two distinct groups by different bubble size, with consis-tent higher pH and ammonia removal rate for micro bubble aeration, under treatment conditions involving liquid depth and aeration rate at constant temperature ( Fig. 2CF). The use of micro bubbles could increase pH by almost 1 unit greater as compared with macro bubble aeration ( Fig. 2E). Ammonia removal rate reached 100% in 6 h in micro bubble aeration, while requiring 12 h in macro bubble aeration ( Fig. 2F). It is obvious that bubble size had the most significant effect on pH elevation and ammonia removal as compared to the liquid depth and airflow rate.

Ammonia removal is mainly controlled by the ammonia diffu-sion through gas film [32]. A decrease in the gas bubble size increases the gasliquid surface area, which in turn controls the amount of free ammonia diffused from water. The gas dispersion has critical importance in determining the performance of a gas liquid system. To keep a maximum level of interfacial area and to enhance the transport phenomena, small bubbles dispersed in uniform form are required.

3.2.3. Airflow rate

Increase of airflow rate accelerates pH elevation and the ammo-nia stripping process ( Fig. 2E and F). The pH increased by approx-imately 0.3 with airflow rate changing from 10 LPM to 35 LPM in both micro bubble and macro bubble aeration, although the effect of airflow rate was more significant in micro bubble aeration. The process of ammonia removal was expedited by the increase of air-flow rate from 10 LPM to 35 LPM in micro bubble aeration, but it was not affected by the variation of airflow rate in macro bubble aeration ( Fig. 2F). In the macro bubble aeration, mass transfer was not enhanced because the surface area was rather reduced by the high airflow rate, which led to gas column formation instead of individual bubbles. Park and Kim studied the kinetics of ammonia stripping and determined the pseudo-first-order rate constant, which showed high initial pH and airflow rate increased the ammonia nitrogen removal rate [33]. However, the issue of decreasing temperature, water evaporation and foaming was noticed when the aeration rate was relatively high. Lei, Sugiura, Feng and Maekawa [21] also observed that the effect of airflow rate was less significant at high flow rate.

Increase of air flow rate decreases the size of gas bubbles dis-persed in the liquid and increases gas holdup, due to increased gas entrainment and gasliquid interfacial area, thus increasing efficiency of ammonia removal as well as mass transfer coefficientQ.-B. Zhao et al. / Chemical Engineering Journal 279 (2015) 313735

Fig. 2. Effect of liquid depth, bubble size, and airflow rate on changes of pH and TAN removal from the AD effluent.

[29,34]. The overall mass transfer resistance for ammonia removal forms mainly in the gas film side because of high dissolution of ammonia in water, leading to overall mass transfer resistance reduction by increasing the airflow rate.

3.3. The chemical equilibrium shifts during ammonia stripping

The chemical equilibrium shifts of carbonate/bicarbonate and ammonia/ammonium were simulated by model ( Fig. 3AD). It was found that DIC dropped one order after 4 h of aeration then decreased slowly until the end of aeration. As the main component of DIC in AD effluent, the concentration of HCO_3 followed a similar

trend to that of DIC but decreased in a faster rate at the beginning. The concentration of CO23_ increased one order and became the dominant species among DIC within 1 h, then followed the exact same trend of decreasing DIC. The CO2 concentration in the outlet gas dropped very fast in the first hour then reached a minimum in 4 h. The initial concentration of CO2 (aq.) was high and the driving force for CO2 to escape from the liquid was also high according to Henrys law. The DIC results shown in Fig. 3A confirmed the fast CO2 release in the first few hours. At the end of aeration DIC was removed by 80%, which indicates aeration is a very effective way to remove CO2 from AD effluent. The CO2 release resulted in an immediate pH elevation while the total concentration of36Q.-B. Zhao et al. / Chemical Engineering Journal 279 (2015) 3137

Fig. 3. Profiles of CO2/HCO_3 /CO23_ and NH3/NH+4 shift during the aeration.

CO23_=HCO_3 and the chemical equilibrium limited further increase of pH.

The rapid decrease of CO2 (aq.) in the first hour allowed chem-ical reactions of Eqs. (1)(3) to proceed towards the right. As OH_ was produced during the shift of Eq. (3) and the initial concentra-tion of HCO_3 was much higher than that of CO23_, Eq. (4) moved towards the left at the beginning of aeration then shifted to the right, which made the CO23_ increase in the first hour and then drop slowly ( Fig. 3A). After CO23_ reached a certain concentration level, H+ primarily resulted from Eqs. (5) and (7). The concentration of both HCO_3 and CO23_ kept decreasing after the first hour.

As the pH in AD effluent is lower than the pKa of ammo-nia/ammonium, ionized ammonium is the main component of TAN. The rapid CO2 release caused a quick pH increase at the begin-ning of the aeration, which allowed the free ammonia concentra-tion to increase during the first hour. Although the concentration of free ammonia increased, it was also noticed that the net increase of free ammonia concentration was less than the net decrease of ammonium concentration, which means the ammonia removal from the liquid started in the first hour of aeration ( Fig. 3C). With pH elevated higher than the pKa of ammonia/ammonium within the first hour of aeration, free ammonia dominated in the liquid phase, then both ammonia and ammonium decreased as aeration kept taking ammonia out of the liquid. The release of ammonia also contributed to a pH drop as ammonia is considered as part of alkalinity.

Ammonia and carbon dioxide may directly react in the liquid phase to form ammonium carbamate (NH2COO_). The release of CO2 promotes Eq. (6) shift to the left, leading to simultaneous release of CO2 and ammonia as well as a pH rise at the beginning

of aeration. Although CO2 could act as a transporting compound which transports ammonia bounded as carbamate, it mainly affects alkalinity, and depletion of CO2 leads to pH elevation [25]. The release of ammonia is favored by low CO2 concentration, which explains that most ammonia release occurred after the majority of original dissolved gaseous CO2 was depleted, pointing to the fact that low CO2 content gas should be used for ammonia stripping.

It was noticed that there are differences between experimental data and model simulation ( Fig. 3D). The model used here describes the concentration CO2=CO23_=HCO_3 at a respective equi-librium state, while the experimental data was obtained from a dynamic aeration process, and the concentration of each species was changing against time. After chemical equilibrium was estab-lished, by the end of aeration process, the difference between experiment and model diminished as the CO2 (aq.) depleted.

3.4. Perspective

A suitable increase in pH is the most critical issue in ammonia stripping process for nitrogen recovery from AD effluent. An effec-tive direct aeration method was successfully applied to elevate pH and recover ammonia simply through aeration without alkali addi-tion, packing material, and tower structure. This strategy is espe-cially suited for high alkalinity AD effluent like digested dairy wastewater. In this study, the TAN removal could reach over 90% within 6 h aeration at a moderate temperature, aeration rate and with micro bubble diffuser.

Fast pH increase by rapid CO2 release led to an instant chemical transformation from ionized ammonium to free ammonia at the beginning of the aeration process. However, the major release ofQ.-B. Zhao et al. / Chemical Engineering Journal 279 (2015) 313737

ammonia was delayed for a couple of hours, indicating desorption of free ammonia from water is rate-limiting step during the aera-tion process. Thus, every effort should be made to facilitate mass transfer and ammonia desorption. Bubble size had the most pre-dominant effect on TAN removal as small bubbles enhanced the mass transfer by increasing the gasliquid surface area as well as KLa. The low liquid level helps keep a high CO2 and ammonia driv-ing force and reduces stripping air pressure. Although overall mass transfer resistance reduced by increase of airflow rate, the effect of airflow rate had less effect on TAN removal at high flow rate. With recovered waste engine heat, increased temperature reduced the ammonia ionization constant (Ka), promoting ionized ammonium shifting to free ammonia, thus shortening the aeration time and lowering the pH requirement.

Although this process is time consuming compared to other physical chemical methods, it presents a robust, simple, cost-efficient and low maintenance way to strip ammonia from digested effluent containing high suspended solids content and thereby control the nutrient levels on their farm and produce prof-itable and exportable fertilizer at the same time. Information con-tained within this study can assist engineers in design and scale-up for optimal control of identified controlling parameters: bubble size, liquid depth/surface area, aeration rate, temperature, and retention time.

4. Conclusion

An effective direct-aeration ammonia stripping method for nitrogen recovery from AD effluent was successfully presented in this study. At moderate temperature (55 LC), the TAN removal could reach over 90% within 6 h aeration with micro bubble dif-fuser. Bubble size was identified as the most important factor with smaller bubble size favoring the process, while airflow rate had the less effect at high flow rates. Concentration profiles of crucial chemical species were satisfactorily predicted with a chemical equilibrium model. Compared to other physical chemical methods, this strategy is robust, simple and cost-efficient, although it requires longer retention time.

Acknowledgments

This research was funded by the following Grant funds: #69-3A75-10-152 from the USDA NRCS Conservation Innovation Grant; #2012-6800219814 from the USDA National Institute of Food and Agriculture; U.S. Environmental Protection Agency Grant number RD-83556701 and the Water Environment Research Foundation, as well as the Washington State University Agricultural Research Center. Thanks are given to two undergradu-ate laboratory assistants, Cynthia Alwine and Alex Dunsmoor.

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