Yield-scaled global warming potential from N2O emissions and CH4 oxidation for almond (Prunus dulcis) irrigated with nitrogen fertilizers on arid land

Agriculture, Ecosystems and Environment 155 (2012) 7–15

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Yield-scaled global warming potential from N2 O emissions and CH4 oxidation for almond (Prunus dulcis) irrigated with nitrogen fertilizers on arid land Daniel L. Schellenberg a,∗ , Maria M. Alsina a , Saiful Muhammad b , Christine M. Stockert a , Michael W. Wolff a , Blake L. Sanden c , Patrick H. Brown b , David R. Smart a a

Department of Viticulture and Enology, University of California, One Shields Avenue, Davis, CA 95616, USA Department of Plant Sciences, University of California, One Shields Avenue, Davis, CA 95616, USA c Kern County Cooperative Extension, University of California, 1031 South Mount Vernon Avenue, Bakersfield, CA 93307, USA b

a r t i c l e

i n f o

Article history: Received 21 October 2011 Received in revised form 16 February 2012 Accepted 11 March 2012 Keywords: Nitrous oxide Methane oxidation Fertilizer source Micro-irrigation Fertigation Perennial crop Arid land Yield-scaled GWP

a b s t r a c t The optimum yield-scaled global warming potential (GWP) of perennial crops on arid land requires effective strategies for irrigation and fertilization. In 2009–2010, N2 O emissions and CH4 oxidation were measured from an almond [Prunus dulcis (Mill.) D.A. Webb] production system irrigated with nitrogen (N) fertilizers. Individual plots were selected within a randomized complete block design with fertilizer treatments of urea ammonium nitrate (UAN) and calcium ammonium nitrate (CAN). Event-related N2 O emissions from irrigation and fertilization were determined for seasonal periods of post-harvest, winter, spring and summer. Peak N2 O emissions in summer occurred within 24 h after fertilization, and were significantly greater from UAN compared to CAN (p < 0.001). Cumulative N2 O emissions from UAN were on average higher than CAN though not significantly different. Air temperature, water-filled pore space (WFPS), soil ammonium (NH4 + ) and soil nitrate (NO3 − ) showed significant positive correlation with N2 O emissions and significant negative correlation was found for the number of days after fertilization (DAF). The percentage of N2 O loss from N fertilizer inputs was 0.23% for CAN and 0.35% for UAN while CH4 oxidation offset 6.0–9.3% of N2 O emissions. Total kernel yield was not significantly different between fertilizer treatments. Yield-scaled GWP for almond from CAN (60.9 kg CO2 eq Mg−1 ) and UAN (91.9 kg CO2 eq Mg−1 ) represent the first report of this metric for a perennial crop. These results outline effective irrigation and fertilization strategies to optimize yield-scaled GWP for almond on arid land. © 2012 Elsevier B.V. All rights reserved.

1. Introduction In the decade of 1999–2009, global agriculture equipped over 31 million ha of cropland with irrigation systems and added 20 million ha of perennial crops (FAO, 2009). During the same period, almond production in California grew 50% from 194,250 ha to 291,375 ha (CASS, 1999, 2009). Such growth could influence California agriculture’s contribution of 8% to statewide greenhouse gas (GHG) emissions, of which more than 50% come from nitrous oxide (N2 O) emissions (Brown et al., 2004). Bouwman and Boumans (2002) estimated 0.9% of nitrogen (N) fertilizer inputs are lost as N2 O while similar values have been adopted for regional scale assessments (IPCC, 2006). Many researchers have proposed an evaluation of global warming potential (GWP) scaled to account for crop yields (Mosier et al., 2006; van Groenigen et al., 2010; Linquist et al., 2012). The intensification of agriculture using irrigation and fertilization warrants an analysis of yield-scaled GWP, particularly in perennial

∗ Corresponding author. Tel.: +1 530 754 7144; fax: +1 530 752 0382. E-mail address: [email protected] (D.L. Schellenberg). 0167-8809/$ – see front matter © 2012 Elsevier B.V. All rights reserved. doi:10.1016/j.agee.2012.03.008

crops, which have not been researched as extensively as annual systems. The microbial processes that regulate N2 O emissions and CH4 oxidation are controlled by water-filled pore space (WFPS), soil ammonium (NH4 + ) and nitrate (NO3 − ) (Davidson and Schimel, 1995). The use of micro-irrigation systems adopted by a majority of California almond growers (Lopus et al., 2010) offers a means to manipulate these factors. Applications of soluble N fertilizer may be split throughout the annual production cycle and choices of N fertilizer from urea, NH4 + and NO3 − sources are widely available. Recent research in maize has shown that urea ammonium nitrate (UAN) had higher N2 O emissions than calcium ammonium nitrate (CAN) in Québec, Canada (Gagnon et al., 2011). Urea increased N2 O emissions compared to polymer-coated urea in Colorado, U.S.A. (Halvorson et al., 2010) and had lower N2 O emissions than anhydrous ammonia in Minnesota, U.S.A. (Venterea et al., 2010). Furthermore, NH4 + -based fertilizers have been shown to both increase and inhibit CH4 oxidation (Bodelier et al., 2000; Sitaula et al., 2000). Strategies to better control N2 O emissions and CH4 oxidation from irrigation and fertilization remain unclear and warrant further investigation.

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The following experiment was conducted to address three major objectives: (1) To identify seasonal patterns of N2 O emissions and CH4 oxidation from irrigation and fertilization; (2) To define soil and climatic parameters that influence N2 O emissions and CH4 oxidation from UAN and CAN fertilizers; and (3) To quantify yieldscaled GWP for almond on arid land in California, U.S.A. 2. Materials and methods 2.1. Site description, climate and management The study was conducted on an almond [Prunus dulcis (Mill.) D.A. Webb] ‘Nonpareil’ orchard interplanted with ‘Monterey’ near Lost Hills in Kern County, California, U.S.A. (N 35◦ 30 37 W 119◦ 40 3 ). Trees were planted in 1999 into 40 cm high berms at a density of 215 trees ha−1 on a Milham sandy loam (Fine-loamy, mixed, superactive, thermic, Typic Haplargids). Each plot of one tree occupied 46.5 m2 and was irrigated with a micro-irrigation system consisting of two static sprinklers (42 L h−1 ) (Fig. 1). Fertilizers were injected and applied with this system, wetting the berm (15.5 m2 ; bulk density 1.1 Mg m−3 ) and the edge (15.5 m2 ; bulk density 1.3 Mg m−3 ). The remaining alleyway space (15.5 m2 ; bulk density 1.5 Mg m−3 ) was maintained without irrigation or cultivation (Fig. 1). Alleyway N2 O and CH4 emissions were used as a baseline since no irrigation water or N fertilizer was applied to this area. Field measurements were carried out year-round with seasonal periods defined as spring (March–May, 2009 and 2010), summer (June–August, 2009 and 2010), post-harvest (September–November, 2009 and 2010) and winter (December 2009–February 2010). No data were collected in winter during the first year (December 2008–February 2009). The experiment has a randomized complete block design with two treatments and five blocks for a total of ten sampled plots. Soluble N fertilizer treatments consisted of UAN (32% N) formulated with 50% ureaN, 25% NH4 + -N and 25% NO3 − -N, and CAN (17% N) formulated with 68% NO3 − -N and 32% NH4 + -N (Yara North America Inc., Tampa, FL). The total annual N fertilizer input of 224 kg N ha−1 was split among the seasonal periods with 45 kg N ha−1 during post-harvest, 45 kg N ha−1 in winter, 67 kg N ha−1 in spring and 67 kg N ha−1 in summer. The fertilizers were injected into the micro-irrigation system over a 3–4 h period during the first half of a 24 h irrigation cycle

in 2009 on February 23rd, April 15th, June 22nd and October 20th in 2009 and on February 22nd, April 7th, June 30th and November 8th in 2010. Irrigation events were scheduled to meet evapotranspiration (ETc ) demand as determined by volumetric water content ( v ) in the root zone using a 503DR neutron probe (Campbell Pacific Nuclear, Concord, CA) and eddy covariance from a surface renewal tower established above the tree canopy (Paw et al., 1995). Irrigation events during summer were scheduled weekly for 24-h sets with occasional 48-h sets to completely recharge soil moisture to 1 m depth. Air temperature and precipitation data were collected from station #146 of the California Irrigation Management and Information System located within 2.5 km of the study site (CIMIS, 2009). Study site climate was characteristic of the South San Joaquin Valley in California U.S.A. During 2009 and 2010, daily average air temperature ranged from 2 to 32 ◦ C and daily average soil temperature ranged from 7 to 32 ◦ C. Rainfall totaled 98 mm in 2009 and 236 mm in 2010 and occurred mainly between the months of November and February. Evapotranspiration demand in 2009 totaled 1563 mm and 1395 mm in 2010 and total irrigation water applied equaled 1442 mm in 2009 and 1212 mm in 2010. The remainder of water to balance ETc demand came from precipitation and soil moisture depletion (Fig. 2). 2.2. Gas sampling and analysis A static chamber technique was used for gas collection (Livingston and Hutchingson, 1995). Gas was extracted at regular intervals (0, 40 and 80 min) from closed chambers using a syringe and was injected into 10 mL evacuated exetainers (Exetainer® , Labco Limited, Buckinghamshire UK). Chambers were cylindrical with a volume of 3.30 L, surface area of 0.03 m2 and height of 11 cm. Each chamber was equipped with a vent to maintain atmospheric pressure, a fan to mix air and a thermocouple to record temperature in the headspace. Gas sampling was conducted to capture both spatial and temporal variability. Three chambers per plot were deployed with one on each of the berm, edge and alleyway spaces. During 1, 2, and 3 days after fertilization (DAF), multiple rounds of gas sampling were conducted at predawn, early morning and late afternoon to account for diurnal variation. On the other days, a single round of gas sampling was conducted in the late afternoon only. Gas sampling continued at 1–7 day intervals after peak N2 O emissions for up to 21 days depending on the seasonal period. Between fertilization events, gas sampling was conducted at 14day intervals. The timing of the 14-day intervals coincided with multiple emission scenarios such as one day after irrigation and precipitation events as well as during dry periods. Samples were analyzed using a gas chromatograph (GC-2014, Shimadzu Scientific Instruments, Columbia, MD) equipped with a 63 Ni electron capture detector for measuring N O gas and a dual 2 flame ionization detector for measuring CH4 gas. Linear regression of chamber GHG concentration over time was used and values were accepted when the correlation coefficient was greater than 95% (R2 > 0.95). The chamber volume, chamber temperature, atmospheric pressure (0.988 atm), soil surface area were used to convert values into GHG emissions (JN2 O-N and JCH4 -C ). Field emissions (JFIELD ) for both N2 O and CH4 were extrapolated to the orchard scale using measurements from the berm, edge and the alleyway: JFIELD = [(JBERM ) × (15.5 m2 tree−1 ) + (JEDGE ) × (15.5 m2 tree−1 )

Fig. 1. Spatial representation of an almond tree plot; the solid square is the tree center, white diamonds are sprinklers, the solid line circles are irrigated area divided by a dashed line into the berm and edge. Area outside the circles represents alleyway space.

+ (JALLEY ) × (15.5 m2 tree−1 )] × (215 trees ha−1 ) × 10−6 where JFIELD is g ha−1 day−1 and JBERM , JEDGE , and JALLEY are ␮g m−2 day−1 .

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Fig. 2. Average daily air and soil temperature, total depth of water from irrigation and rainfall, and evapotranspiration (ETc ) during seasonal periods in 2009 and 2010.

Direct emissions (JDIRECT ) are reported herein and were calculated by taking the difference between JFIELD and baseline emissions from the non-irrigated and uncultivated alleyway: JDIRECT = JFIELD − [(JALLEY ) × (46.5 m2 tree−1 ) × (215 trees ha−1 )] × 10−6 where JDIRECT and JFIELD are g ha−1 day−1 and JALLEY is ␮g m−2 day−1 Direct emissions were partitioned into seasonal periods and integrated using the trapezoid rule to generate cumulative N2 O emissions and CH4 oxidation for spring 2009, summer 2009, postharvest 2009, winter 2009–2010, spring 2010, summer 2010 and post-harvest 2010. The annual production cycle for 2010 almond crop consisted of seasonal periods from post-harvest 2009 until summer 2010.

(Sercon Ltd., Cheshire, UK). Soil pH was measured from soil slurry mixed 1:1 with de-ionized distilled water. Concurrent with each day of gas sampling, soil samples from each treatment plot were extracted from both the berm and the edge using a 2.5 cm diameter push core tool to 30 cm depth. Soil was oven-dried at 105 ◦ C for determination of gravimetric water content, and water-filled pore space (WFPS) was calculated using bulk density from either the berm (1.1 Mg m−3 ) or the edge (1.3 Mg m−3 ). A separate aliquot of 20 g fresh weight soil was extracted in 80 mL 2M KCl and analyzed colorimetrically following Kempers and Kok (1989) for NH4 + and Doane and Horwath (2003) for NO3 − . Values for soil mineral N content were the sum of KCl-extractable NH4 + -N and NO3 − -N. Soil temperature was recorded using a temperature probe at 15 cm depth. Averaged values of WFPS, soil NH4 + -N and NO3 − -N from the berm and edge are reported herein.

2.3. Soil sampling and analyses 2.4. Almond harvest Soils were collected for analysis of physical and chemical properties at the termination of the experiment (Table 1). Soil particle size analysis was determined using the hydrometer method (Gee and Bauder, 1986). Soils for analysis of total organic carbon and total nitrogen were ball-milled, packed into silver capsules, acid-fumigated for carbonate removal (Harris et al., 2001) and combusted at 1000 ◦ C by a PDZ Europa ANCA-GSL elemental analyzer

Table 1 Soil physical and chemical properties from plots within a randomized complete block design amended with fertilizer treatments CAN and UAN. Soils were harvested under gas sampling sites at the termination of the experiment.

Bulk density (Mg m−3 ) Sand (%) Silt (%) Clay (%) Total organic carbon (g kg−1 ) Total nitrogen (g kg−1 ) pH

CAN

UAN

1.20 64.2 16.5 19.3 4.01 0.47 7.56

1.20 63.8 17.6 18.6 4.10 0.48 7.34

All trees were shaken on August 27th 2010 at full kernel maturity to allow for partial drying. Drying of fruits on the orchard floor continued for 10 days to achieve final moisture content between 5 and 7%. The fruit was then swept into windrows for pickup by the harvester. The almond fruit consists of three separate parts – the hull, shell and kernel. Sub-samples were collected from the windrows just before pickup to determine mass fraction of the kernel from total almond yield, which averaged 28%. 2.5. Yield-scaled GWP Direct N2 O emissions and CH4 oxidation were converted to GWP units of carbon dioxide equivalents (CO2 eq) within a 100-year horizon by multiplying by a radiative forcing potential equivalent to CO2 of 298 for N2 O and 25 for CH4 (IPCC, 2001). Cumulative GHG emissions (CO2 eq ha−1 year−1 ) for the annual production cycle from 2009 to 2010 divided by total almond kernel yield (Mg ha−1 year−1 ) equaled yield-scaled GWP for this perennial system (CO2 eq Mg−1 ).

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2.6. Statistical analysis Log-normally transformed data for N2 O emissions and data for CH4 oxidation were analyzed for significant correlation with soil and climatic parameters by single step-wise and multiple linear regression in R (R Project, http://r-project.org). Cumulative N2 O emissions and CH4 oxidation for seasonal periods and percentage N2 O lost from N fertilizer input, GWP, kernel yield and yield-scaled GWP for the annual production cycle were analyzed for significance (p < 0.05) between treatments using PROC GLM for ANOVA (SAS Cary, NC). Figures were constructed using Sigma Plot 12.0 (San Jose, CA).

3. Results and discussion 3.1. Nitrous oxide emissions 3.1.1. Peak seasonal emissions During seasonal fertilization events, N2 O emissions averaged across fertilizer treatments followed consistent observable patterns in both years. Peak emissions were observed at 3 DAF in winter 2010 (12.1 g N2 O-N ha−1 day−1 ). In spring of 2009 (15.2 g N2 O-N ha−1 day−1 ), spring 2010 (15.6 g N2 O-N ha−1 day−1 ), post harvest 2009 (9.6 g N2 O-N ha−1 day−1 ) and post harvest 2010 (3.2 g N2 O-N ha−1 day−1 ) peak emissions were observed at 2 DAF. Peak emissions occurred within 24 h after fertilization in summer 2009 (25.5 g N2 O-N ha−1 day−1 ) and summer 2010 (23.9 g N2 ON ha−1 day−1 ) (Fig. 3). The amount of time elapsed to return to pre-fertilization levels varied by season and depended on the timing of irrigation after fertilization (see Section 3.1.4). Earlier peak emissions, such as in summer, resulted in a more rapid return to pre-fertilization levels while a slower pace to peak emissions, as in winter, prolonged the emissions after a fertilization event (Fig. 3). The average daily emissions from the uncultivated alleyway were

negligible compared to the irrigated and fertilized berm and edge (<0.1 g N2 O-N ha−1 day−1 ). Soil conditions following irrigation and fertilization were characterized by increased WFPS while increased soil mineral N was restricted to fertilization events (Fig. 4). In the bulk soil, WFPS was never observed to exceed 50% in this well-drained sandy loam. Peak N2 O emissions were consistent with WFPS greater than 30% and soil mineral N greater than 30 mg N kg−1 soil, except in postharvest 2009 and winter 2010 when soil mineral N was closer to 20 mg N kg−1 soil (Fig. 4). A decline in N2 O emissions generally corresponded with a decrease in WFPS to less than 20% and soil mineral N less than 10 mg N kg−1 soil. Similar patterns were evident in other studies of both annual and perennial crops. In a California vineyard, peak N2 O emissions were observed immediately after spring fertigation with UAN (Garland et al., 2011). In row crop production in California and China, N fertilizer side-dressed or in a band application resulted in peak N2 O emissions after subsequent irrigations (Lee et al., 2009; Liu et al., 2010). To the contrary, sugarcane production in Australia fertilized with liquid urea and granular NH4 NO3 led to peak N2 O emissions many weeks after application (Allen et al., 2010), perhaps because of soil NO3 − accumulation due to NO3 − discrimination by the sugarcane crop (Robinson et al., 2011). In our study, soluble N fertilizer was delivered through a micro-irrigation system with 40 mm H2 O. Consequently, the N was readily available for microbial processes that regulate N2 O emissions. Throughout the experiment both fertilizer sources reached peak emissions at the same DAF, varying by season. This observation suggests that combined effects of parameters such as WFPS and temperature influence these seasonal patterns. Field conditions after fertilization in this study were similar to observations by Conen et al. (2000) who reported an increase in N2 O emissions with increasing soil temperature when WFPS was constant and soil mineral N was not limiting. The results from this study show higher soil temperatures led to both earlier and greater peak N2 O emissions.

Fig. 3. Seasonal patterns of N2 O emissions and CH4 oxidation averaged across fertilizer treatments of urea ammonium nitrate (UAN) and calcium ammonium nitrate (CAN) during spring, summer, post-harvest and winter in 2009 and 2010. Data points represent mean values with standard error bars (n = 10).

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Fig. 4. Seasonal patterns of water-filled pore space (WFPS) and soil mineral N (NH4 + -N + NO3 − -N) averaged across fertilizer treatments of urea ammonium nitrate (UAN) and calcium ammonium nitrate (CAN) during spring, summer, post-harvest and winter in 2009 and 2010. Data points represent mean values with standard error bars (n = 10).

Thus, the magnitude and time of the N2 O emission peak after fertilization were seasonally dependent and may be predicted by soil temperature. These conclusions may only be applicable to perennial systems on arid land, since past research has demonstrated that the temperature optima for nitrification and denitrification are ecosystem-specific (Powlson et al., 1988; Stark and Firestone, 1996). 3.1.2. Soil and climatic parameters Multiple linear regression yielded significant coefficients for soil and climatic parameters in both fertilizer treatments (p < 0.05). Consistent with observations outlined above, WFPS, soil NH4 + -N and NO3 − -N, and air temperature showed positive correlations with N2 O emissions, while DAF had a negative correlation. The positive correlation with soil NO3 − -N and N2 O emissions was only significant for the UAN treatment. No significant correlation

between soil temperature and N2 O emissions was observed (data not shown). The statistical model generated less residual error for UAN than CAN and N2 O emissions from both treatments were well-correlated with coefficients for soil and climatic parameters (Table 2). Coefficients for WFPS, soil NH4 + -N and DAF when UAN was the N source showed a stronger correlation with N2 O emissions than when CAN was the N source. Surprisingly, given the higher concentration of NO3 − -N in the CAN fertilizer formulation, coefficients for soil NO3 − -N were four times greater for UAN and insignificant for CAN. Air temperature had the same influence on emissions for both UAN and CAN (Table 2). Under both treatments, soil NH4 + -N had a greater explanatory value for N2 O emissions than did soil NO3 − -N. The same phenomenon was apparent in studies in a California vineyard (Garland et al., 2011) and in a sugarcane field in Australia (Allen et al., 2010). These results may suggest a greater

Table 2 Multivariate correlation coefficients (ˇ) and statistical probabilities (p) for log-normally transformed N2 O emissions (JN2 O ) and CH4 oxidation (JCH4 ) under fertilizer treatments of urea ammonium nitrate (UAN) and calcium ammonium nitrate (CAN) pooled for all seasonal periods during 2009 and 2010. Parameters of water-filled pore space (WFPS), soil ammonium (NH4 + -N), soil nitrate (NO3 − -N), air temperature and days after fertilization (DAF) are fitted to a multiple linear regression equationa with generation of residual error (ε) and R2 values. Log normal (g N2 O-N ha−1 day−1 ) CAN

WFPS (%) NH4 + -N (mg kg−1 ) NO3 − -N (mg kg−1 ) Air temp (◦ C) DAF (day) E R2

g CH4 -C ha−1 day−1 UAN

CAN

UAN

B

p

ˇ

p

ˇ

p

B

p

0.023 0.031 0.005 0.068 −0.017

0.002 <0.001 0.248 <0.001 <0.001

0.031 0.039 0.021 0.068 −0.023

<0.001 <0.001 <0.001 <0.001 <0.001

−0.010 −0.012 0.003 0.041 0.002

0.131 0.101 0.394 0.001 0.174

−0.008 −0.009 0.003 0.031 −0.001

0.193 0.204 0.378 0.014 0.545

0.860 0.502

0.822 0.668

Values in bold are significant (p < 0.05). a JN2 O-N or JCH4 -C = ˇ1 (WFPS) + ˇ2 (NH4 + -N) + ˇ3 (NO3 − -N) + ˇ4 (Air Temp) + ˇ5 (DAF) + ε.

0.708 0.200

0.708 0.132

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Fig. 5. Week-long N2 O emissions in summer from fertilization with 67 kg N ha−1 on Day 0 with soil drying in summer 2009 and irrigation at 5 DAF in summer 2010. Peak N2 O emissions within 24 h after fertilization from UAN** were significantly greater than from CAN* (p < 0.001). Data points represent mean values with standard error bars (n = 5).

role of nitrification for N2 O emissions from perennial and sugarcane crops.

3.1.3. Effects of fertilizer N source Nitrogen source had a significant effect on peak N2 O emissions in this perennial system after fertilization in summer. Peak N2 O emissions from UAN in 2009 (37.1 g N2 O-N ha−1 day−1 ) and 2010 (36.6 g N2 O-N ha−1 day−1 ) were significantly greater (p < 0.001) than peak N2 O emissions from CAN in 2009 (13.9 g N2 ON ha−1 day−1 ) and 2010 (11.2 g N2 O-N ha−1 day−1 ) (Fig. 5). As soils dried during the subsequent days following the fertilization event, N2 O emissions declined rapidly, and any differences observed between UAN and CAN were no longer significant (p > 0.05). Other studies demonstrated N source effects on N2 O emissions. Gagnon et al. (2011) reported higher N2 O emissions from UAN compared to CAN under maize production on a clay soil in Québec, Canada. Maize production research on a clay loam soil in Colorado showed higher N2 O emissions from urea when compared to a polymer-coated urea designed to inhibit nitrification (Halvorson et al., 2010). In laboratory incubations on a clay loam at 60% WFPS, Liu et al. (2007) observed greater 15 N2 O flux from soils amended with 15 NH4 + than with 15 NO3 − . In general, numerous investigations suggest greater N2 O emissions when urea and NH4 + are part of the N source and therefore call greater attention to the role of nitrification in agricultural systems. Higher peak N2 O emissions from UAN compared to CAN may also be caused by the production of NO3 − from nitrification, which is subsequently available for denitrification. The microbial

processes associated with denitrification are fostered by anaerobic conditions. The lack of available oxygen (O2 ), as a consequence of the low solubility of O2 in water combined with its rapid consumption by root and microbial respiration, enhances denitrification activity when WFPS is greater than 60% (Linn and Doran, 1984; Khalil et al., 2004). The WFPS was consistently less than 60% in this perennial system, even following irrigation, which suggests low rates of denitrification and may offer an explanation for overall low cumulative N2 O emissions (see Section 3.1.5). Under UAN fertilization greater NH4 + inputs may be held in the upper soil horizons and be more available for microbial processes. In contrast, greater NO3 − inputs from CAN may move to lower soil horizons where there are fewer denitrifiers, which may explain lower N2 O emissions (Smart et al., 2011).

3.1.4. Effects of irrigation In summer 2009, fertilization was followed by a week-long dry period when N2 O emissions declined to pre-fertilization levels within 7 DAF. During summer 2010, drier conditions led to irrigation at 5 DAF (Fig. 5). Following irrigation after fertilization in summer 2010, N2 O emissions from UAN (14.3 g N2 O-N ha−1 day−1 ) and CAN (10.4 g N2 O-N ha−1 day−1 ) were lower than the peak levels observed within 24 h after fertilization (see Section 3.1.1). This observation verifies that higher N2 O emissions occur from fertilization with water compared to irrigation water only. From irrigated cotton on arid land in Uzbekistan, Scheer et al. (2008a) reported that N2 O emissions from irrigation were substantially lower than N2 O emissions from fertilization. Similar findings were evident in

Table 3 Cumulative N2 O emissions (kg N2 O-N ha−1 ) and percentage N2 O loss from N fertilizer input (%) during seasonal periods in 2009 and 2010; post-harvest and winter receive 45 kg N ha−1 and spring and summer receive 67 kg N ha−1 . Data represent mean values plus or minus standard error (n = 5). Winter

CAN UAN CAN UAN

CAN UAN CAN UAN a

N2 O emissions (kg N2 O-N ha−1 )a 2009 – – 2010 0.09 ± 0.02 0.11 ± 0.03

Spring

Summer

Post-harvest

0.15 ± 0.05 0.20 ± 0.06

0.11 ± 0.04 0.26 ± 0.08

0.08 ± 0.02 0.12 ± 0.02

0.19 ± 0.06 0.27 ± 0.08

0.16 ± 0.05 0.28 ± 0.07

0.05 ± 0.01 0.11 ± 0.02

0.16 ± 0.06 0.39 ± 0.12

0.17 ± 0.04 0.26 ± 0.04

0.24 ± 0.07 0.42 ± 0.10

0.11 ± 0.02 0.25 ± 0.04

Percentage N2 O loss from N fertilizer input (%)a 2009 – 0.22 ± 0.07 – 0.30 ± 0.09 2010 0.20 ± 0.04 0.28 ± 0.09 0.25 ± 0.07 0.40 ± 0.12

No significant differences were observed between fertilizer treatments (p > 0.05).

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our study, where daily N2 O emissions after irrigation of less than 5.0 g N2 O-N ha−1 day−1 were measured multiple times in between fertilization events (Fig. 3). Our results demonstrate that greater emphasis on fertilization strategies to reduce N2 O emissions is evident. However, timing irrigation more than 7 DAF may be an effective management strategy for minor reductions in N2 O emissions from this perennial system. 3.1.5. Cumulative seasonal emissions Cumulative N2 O emissions ranged from 0.11 to 0.42% of the total N fertilizer input, varying by seasonal periods (Table 3). These percentages are substantially lower than those of Bouwman and Boumans (2002) who indicated that on average about 0.9% of N fertilizer is lost as N2 O. During the seasonal periods of winter and post-harvest, a lower N fertilizer input per event (45 kg N ha−1 ) was applied, while in spring and summer a higher N fertilizer input per event (67 kg N ha−1 ) was applied. Cumulative N2 O emissions were consistently greater from UAN compared to CAN across all seasonal periods and fertilization events, although the differences were not significant (p > 0.05). There were notable patterns in the percentage of N2 O loss from N fertilizer input, varying by seasonal period, but the quantity of N fertilizer input was different as well. The percentage loss was greater during spring and summer as compared to post-harvest and winter except during summer 2009 (0.16%) and post-harvest 2009 (0.17%) for CAN (Table 3). These results suggest that higher N fertilizer inputs during warmer seasons increase the percentage of N2 O loss. Split applications of N fertilizer on cotton in Uzbekistan resulted in greater N2 O emissions during warmer seasons (Scheer et al., 2008a,b) while lower N2 O emissions were observed in summer compared to spring fertilization from California maize production (Lee et al., 2009). The percentage of N2 O loss from application of N fertilizer split between spring and summer was 2.95% versus 6.70% from a single spring application in an Australian sugarcane production system (Allen et al., 2010). A strategy to split N fertilization between seasonal periods is less likely with many annual crops that receive single applications of N fertilizer (Halvorson et al., 2010; Liu et al., 2010; Venterea et al., 2010) and may only be suitable for perennial crops fertilized with micro-irrigation systems. 3.2. CH4 oxidation 3.2.1. Patterns and controls During the study period, steady CH4 oxidation indicated that this perennial system was a net CH4 sink. Occasional positive daily emissions were observed and shared no discernible pattern with N2 O emissions (Fig. 3). Intensive management of agricultural crops has been shown to stimulate both N2 O emissions and CH4 oxidation (Robertson et al., 2000). Multiple production systems have shown CH4 oxidation throughout the growing season with spikes in CH4 emissions during rainfall or irrigation events (Delgado et al., 1996; Weier, 1999; Toma et al., 2011). Average daily CH4 emissions (<0.1 g CH4 -C ha−1 day−1 ) from the uncultivated alleyway were low for this perennial system. Fertilization through the micro-irrigation system may have stimulated CH4 oxidation compared to the uncultivated alleyway. However, these differences were not significantly different (p > 0.05) and may be the result of soil compaction. Higher bulk density in the alleyway may have inhibited CH4 oxidation (Sitaula et al., 2000). Methane oxidation from UAN and CAN was less correlated with soil and climatic parameters than were N2 O emissions (see Section 3.1.2) (Table 2). Air temperature was the only significant coefficient for CH4 oxidation and showed a positive correlation. For both treatments, WFPS and soil NH4 + -N were negatively correlated with CH4 oxidation while soil NO3 − -N demonstrated a positive correlation. Previous research has shown that NH4 + fertilizers inhibit CH4

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Table 4 Cumulative CH4 oxidation (g CH4 -C ha−1 ) during seasonal periods during 2009 and 2010; post-harvest and winter receive 45 kg N ha−1 and spring and summer receive 67 kg N ha−1 . Data represent mean values plus or minus standard error (n = 5). Winter

CAN UAN CAN UAN

Spring

CH4 oxidation (g CH4 -C ha−1 )a 2009 – 23.3 ± 7.06 – 25.3 ± 5.32 2010 44.1 ± 9.20 62.2 ± 9.61 68.9 ± 20.1 49.5 ± 12.4

Summer

Post-harvest

29.6 ± 12.2 64.2 ± 9.80

71.3 ± 34.6 81.0 ± 18.6

68.5 ± 18.7 41.7 ± 20.5

74.7 ± 16.9 59.3 ± 32.1

a No significant differences were observed between fertilizer treatments (p > 0.05).

oxidation in upland agricultural soils (Sitaula et al., 2000). However, the regression coefficients for these soil parameters in this study were not significant (p > 0.05). The statistical model generated the same residual error for UAN and CAN. 3.2.2. Cumulative seasonal effects Cumulative CH4 oxidation ranged from 23.3 to 81.0 g CH4 -C ha−1 per seasonal period (Table 4). These results represent an offset of GHG emissions from this perennial system. As with N2 O emissions, there were no significant differences between the seasons with the same N fertilizer input (p > 0.05). However, the amount of CH4 oxidation was greater from seasons following the lower N fertilizer input. Exceptions were observed in summer 2009 from UAN (64.2 g CH4 -C ha−1 ) and in summer 2010 from CAN (68.5 g CH4 C ha−1 ) where the amount of CH4 oxidation was the highest from higher N fertilizer inputs (Table 4). These results suggest CH4 oxidation may be inhibited by higher N fertilizer inputs in spite of the fact that there was no significant correlation with soil NH4 + (Table 3). Research has demonstrated lower CH4 oxidation from N fertilization (Sitaula et al., 2000) and no effect on CH4 oxidation from single and split applications of N fertilizer (Delgado et al., 1996; Weier, 1999). 3.3. Annual production cycle 3.3.1. Global warming potential From this study, several key findings emerged that elucidate the GWP for this perennial system. First, the percentage of N2 O Table 5 Annual production cycle of fertilizer N inputs, N2 O emissions, percentage N2 O loss from N fertilizer input, CH4 oxidation, global warming potential (GWP), almond kernel yield and yield-scaled GWP from seasonal periods of post-harvest 2009, winter 2009–2010, spring 2010, and summer 2010. Data represent mean values plus or minus standard error (n = 5). CANa Nitrous oxide emissions (kg N2 O-N ha−1 year−1 ) Total N fertilizer inputs (kg N ha−1 year−1 ) Percentage N2 O loss from N fertilizer input (%) Nitrous oxide emissions (kg CO2 eq ha−1 year−1 ) Methane oxidation (kg CO2 eq ha−1 year−1 ) Global warming potential (kg CO2 eq ha−1 year−1 ) Total almond kernel yield (Mg ha−1 year−1 ) Yield-scaled GWP (kg CO2 eq Mg−1 )

0.53 ± 0.11 224 0.23 ± 0.05 248 ± 50.7 21.5 ± 5.11 227 ± 54.5 3.81 ± 0.19 60.9 ± 15.5

UAN 0.80 ± 0.19 224 0.35 ± 0.08 375 ± 88.6 21.0 ± 4.55 354 ± 91.1 3.89 ± 0.17 91.9 ± 24.6

a No significant differences were observed between fertilizer treatments (p > 0.05).

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lost from the application of 224 kg N ha−1 was 0.35% from UAN and 0.23% from CAN (Table 5), which was substantially lower than a broad range of agricultural systems (Bouwman and Boumans, 2002). The low N2 O emissions observed may be related to the use of strategies for irrigation and fertilization that target water and N fertilizer with tree demand. A second finding of significance was steady CH4 oxidation across seasons and N fertilizer treatments. This sink resulted in an offset of N2 O emissions of 6.0% for UAN and 9.3% for CAN during the annual production cycle from post-harvest 2009 to summer 2010 (Table 5). The GWP from N2 O emissions and CH4 oxidation for UAN (375 kg CO2 eq ha−1 year−1 ) and CAN (248 kg CO2 eq ha−1 year−1 ) were an order of magnitude lower than an annual crop rotation in an arid climate (Scheer et al., 2008b). The use of micro-irrigation systems combined with split N fertilizer applications that coincide with tree demand fit previously described strategies to lower the GWP of agricultural systems (Freney, 1997; Verma et al., 2006). In California, these strategies are scalable due to the adoption of micro-irrigation systems and the use of them to deliver soluble N fertilizers by a majority of almond growers (Lopus et al., 2010). Chemical tools such as urease or nitrification inhibitors could reduce the greater GWP observed in the UAN treatment (Bremner, 1997; Halvorson et al., 2010). Furthermore, the opportunity to deliver soluble N fertilizer with higher frequency at lower concentrations may decrease GWP if soil mineral N remains below levels optimum for N2 O production and inhibition of CH4 oxidation. 3.3.2. Yield-scaled GWP A recent agronomic assessment of major grain crops reported the average yield-scaled GWP from 22 studies of wheat (166 kg CO2 eq Mg−1 ) and 19 studies of maize (185 kg CO2 eq Mg−1 ) (Linquist et al., 2012). This comprehensive review provides a valuable metric for other agricultural systems where GWP in relation to crop yields is essential for the implementation of effective management strategies. The almond kernel yield in this study was not significantly different between fertilizer treatments (Table 5). As a result, the yield-scaled GWP from UAN (91.9 kg CO2 eq Mg−1 ) was greater than CAN (60.9 kg CO2 eq Mg−1 ) due to higher N2 O emissions generated during the annual production cycle. To our knowledge, this study reports the first account of yield-scaled GWP for a perennial crop. The lowest yield-scaled GWP may not ensure maximum yield. At the fertilization rate of 224 kg N ha−1 in this study, almond kernel yield was 94% and 87% of maximum yield for UAN and CAN, respectively (Muhammad and Brown, unpublished data). Under maximum yield conditions of 308 kg N ha−1 , we would expect yield-scaled GWP to be approximately 18% greater, assuming the same percentage N2 O loss from N fertilizer input and proportion of N2 O emissions offset by CH4 oxidation (Table 5). Similar results were reported for wheat (102 kg CO2 eq Mg−1 ) and maize (140 kg CO2 eq Mg−1 ) where the lowest yield-scaled GWP was achieved at 92% of maximum yield (Linquist et al., 2012). This 11-year old orchard may expect higher yields in the future, while alterations to irrigation and fertilization strategies may further reduce N2 O emissions and stimulate CH4 oxidation to optimize yield-scaled GWP. 4. Conclusions In our study, soluble N fertilizer delivered through a microirrigation system targeted to tree demand resulted in low N2 O emissions. Adjusting for steady CH4 oxidation, we found a relatively low yield-scaled GWP compared to wheat and maize. Our results demonstrate that timing irrigation at more than 7 DAF may lead to minor reductions in N2 O emissions. Peak N2 O emissions

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