Introduction
Improved resource-use efficiencies are pivotal components of sustainable
agriculture that meets human needs and protects natural resources (Spiertz,
2010). Several strategies have been proposed to improve the efficiency of
intensive irrigated systems, where nitrate (NO3-) leaching losses
are of major concern, during both cash crop and winter fallow periods
(Quemada et al., 2013). In this sense, replacing winter intercrop
fallow with cover crops (CCs) has been reported to decrease NO3-
leaching via retention of post-harvest surplus inorganic nitrogen (N)
(Wagner-Riddle and Thurtell, 1998), consequently improving N use efficiency
of the cropping system (Gabriel and Quemada, 2011). Furthermore, the use of
CCs as green manure for the subsequent cash crop may further increase soil
fertility and N use efficiency (Tonitto et al., 2006; Veenstra et
al., 2007) through slow release of N and other nutrients from the
crop residues, leading to a saving in synthetic fertilizer.
From an environmental point of view, N fertilization is closely related to
the production and emission of nitrous oxide (N2O) (Davidson and Kanter,
2014), a greenhouse gas (GHG) with a molecular global warming potential
ca. 300 times that of carbon dioxide (CO2) (IPCC, 2007). Nitrous oxide
released from agricultural soils is mainly generated by nitrification and
denitrification processes, which are influenced by several soil variables
(Firestone and Davidson, 1989). Thereby, modifying these parameters through
agricultural management practices (e.g., fertilization, crop rotation,
tillage or irrigation) aiming to optimize N inputs can lead to strategies for
reducing the emission of (N2O) (Ussiri and Lal, 2013). In order to
identify the most effective GHG mitigation strategies, side effects of
methane (CH4) uptake and CO2 emission (i.e., respiration) from
soils, which are also influenced by agricultural practices (Snyder et al.,
2009), need to be considered.
To date, the available information linking GHG emission and maize–winter CC
rotation in the scientific literature is scarce. The most important knowledge
gaps include effects of plant species selection and CC residue management
(i.e., retention, incorporation or removal) (Basche et al., 2014). Cover crop
species may affect N2O emissions in contrasting ways by influencing
abiotic and biotic soil factors. These factors include mineral N availability
in soil and the availability of carbon (C) sources for the denitrifier
bacterial communities, soil pH, soil structure and microbial community
composition (Abalos et al., 2014). For example, nonlegume CCs such as winter
cereals could contribute to a reduction of N2O emissions due to their
deep roots, which allow them to extract soil N more efficiently than legumes
(Kallenbach et al., 2010). Conversely, it has been suggested that the higher
C : N ratio of their residues as compared to those of legumes may provide
energy (C) for denitrifiers, thereby leading to higher N2O losses in the
presence of mineral N-NO3- from fertilizers (Sarkodie-Addo et al.,
2003). In this sense, the presence of cereal residues can increase the
abundance of denitrifying microorganisms (Gao et al., 2016), thus enhancing
denitrification losses when soil conditions are favorable (e.g., high
NO3- availability and soil moisture after rainfall or irrigation
events, particularly in fine-textured soils) (Stehfest and Bouwman 2006;
Baral et al., 2016). Furthermore, winter CCs can also abate indirect gaseous
N losses through the reduction of leaching and subsequent emissions from
water resources (Feyereisen et al., 2006). Thus, the estimated N2O
mitigation potential for winter CCs ranges from 0.2 to
1.1 kg N2O ha-1 yr-1 according to Ussiri and Lal (2013).
In a CC–maize rotation system, mineral fertilizer application to the cash
crop could have an important effect on N use efficiency and N losses from
the agro-ecosystem. Different methods for calculating the N application rate
(e.g., conventional or integrated) can be employed by farmers, affecting the
amount of synthetic N applied to soil and the overall effect of CCs on
N2O fluxes. Integrated soil fertility management (ISFM) (Kimani et al.,
2003) provides an opportunity to optimize the use of available resources,
thereby reducing pollution and costs from overuse of N fertilizers
(conventional management). ISFM involves the use of inorganic fertilizers
and organic inputs, such as green manure, and aims to maximize agronomic
efficiency (Vanlauwe et al., 2011). When applying this technique to a
CC–maize crop rotation, the N fertilization rate for maize is calculated taking
into account the background soil mineral N and the expected available N from
mineralization of CC residues, which depends on residue composition.
Differences in soil mineral N during the cash crop phase may be
significantly reduced if ISFM practices are employed, affecting the GHG
balance of the CC–cash crop cropping system.
Only one study has investigated the effect of CCs on N2O emissions in
Mediterranean cropping systems (Sanz-Cobena et al., 2014). These authors
found an effect of CC species on N2O emissions during the intercrop
period. After 4 years of CC (vetch, barley or rape)–maize rotation, vetch
was the only CC species that significantly enhanced N2O losses compared
to fallow, mainly due to its capacity to fix atmospheric N2 and
because of higher N surplus from the previous cropping phases in these
plots. In this study a conventional fertilization (same N synthetic rate for
all treatments) was applied during the maize phase; how ISFM practices may
affect these findings remains unknown. Moreover, the relative contribution
of mineral N fertilizer, CC residues and/or soil mineral N to N2O
losses during the cash crop has not been assessed yet. In this sense, stable
isotope analysis (i.e., 15N) represents a way to identify the source and
the dominant processes involved in N2O production (Arah, 1997). Stable
isotope techniques have been used in field studies evaluating N leaching
and/or plant recovery in systems with cover crops (Bergström et al.,
2001; Gabriel and Quemada, 2011; Gabriel et al., 2016). Furthermore, some
laboratory studies have evaluated the effect of different crop residues on
N2O losses using 15N techniques (Baggs et al., 2003; Li et al.,
2016), but to date, no previous studies have evaluated the relative
contribution of cover crops (which include the aboveground biomass and the
decomposition of root biomass) and N synthetic fertilizers to N2O
emissions under field conditions. A comprehensive understanding of the
N2O biochemical production pathways and nutrient sources is crucial for
the development of effective mitigation strategies.
The objective of this study was to evaluate the effect of two different CC
species (barley and vetch) and fallow on GHG emissions during the CC period
and during the following maize cash crop period in an ISFM system. An
additional objective was to study the contribution of the synthetic
fertilizer and other N sources to N2O emissions using 15N-labeled
fertilizer. We hypothesized that (1) the presence of CCs instead of fallow
would affect N2O losses, leading to higher emissions in the case of the
legume CC (vetch) in accordance with the studies of Basche et al. (2014) and
Sanz-Cobena et al. (2014) and (2) in spite of the ISFM during the maize
period, which theoretically would lead to similar soil N availability for
all plots, the distinct composition of the CC residues would affect N2O
emissions. In order to test these hypotheses, a field experiment was carried
out using the same management system for 8 years, measuring GHGs during the
8th year. To gain a better understanding of the effect of the
management practices tested on the overall GHG budget of a cropping system,
CH4, CO2 and yield-scaled N2O emissions were also
analyzed during the experimental period. The relative contribution of each N
source (synthetic fertilizer or soil endogenous N, including N mineralized
from the CCs) to N2O emissions was also evaluated by 15N-labeled
ammonium nitrate in a parallel experiment.
Materials and methods
Site characteristics
The study was conducted at “La Chimenea” field station (40∘03′ N, 03∘ 31′ W; 550 m a.s.l.), located in the
central Tajo River basin near Aranjuez (Madrid, Spain), where an experiment
involving cover-cropping systems and conservation tillage has been carried
out since 2006. Soil at the field site is a silty clay loam (Typic
Calcixerept; Soil Survey Staff, 2014). Some of the physicochemical
properties of the top 0–10 cm soil layer, as measured by conventional
methods, were as follows: pHH2O, 8.16; total organic C, 19.0 g kg-1;
CaCO3, 198 g kg-1; clay, 25 %; silt, 49 %; and sand, 26 %.
Bulk density of the topsoil layer determined in intact core samples
(Grossman and Reinsch, 2002) was 1.46 g cm-3. Average ammonium
(NH4+) content at the beginning of the experiment was 0.42 ± 0.2 mg N kg soil-1
(without differences between treatments). Nitrate
concentrations were 1.5 ± 0.2 mg N kg soil-1 in fallow and barley
and 0.9 ± 0.1 mg N kg soil-1 in vetch. Initial dissolved organic C
(DOC) contents were 56.0 ± 7 mg C kg soil-1 in vetch and fallow
and 68.8 ± 5 mg C kg soil-1 in barley. The area has a
Mediterranean semiarid climate, with a mean annual air temperature of
14 ∘C. The coldest month is January, with a mean temperature of
6 ∘C, and the hottest month is August, with a mean temperature of
24 ∘C. During the last 30 years, the mean annual precipitation
has been approximately 350 mm (17 mm from July to August and 131 mm from
September to November).
Hourly rainfall and air temperature data were obtained from a meteorological
station located at the field site (CR10X, Campbell Scientific Ltd., Shepshed,
UK). A temperature probe inserted 10 cm into the soil was used to measure
soil temperature. Mean hourly temperature data were stored on a data logger.
Experimental design and agronomic management
Twelve plots (12 m × 12 m) were randomly distributed in four
replications of three cover-cropping treatments, including a cereal and a
legume: (1) barley (B) (Hordeum vulgare L. `Vanessa'), (2) vetch
(V) (Vicia sativa L. `Vereda'), and (3) traditional winter fallow
(F). Cover crop seeds were broadcast by hand over the stubble of the
previous crop and covered with a shallow cultivator (5 cm depth) on 10 October
2013, at a rate of 180 and 150 kg ha-1 for B and V,
respectively. The cover-cropping phase finished on 14 March 2014 following
local practices, with an application of glyphosate (N-phosphonomethyl
glycine) at a rate of 0.7 kg a.e. ha-1. Even though the safe use of
glyphosate has been under discussion for many years (Chang and Delzell,
2016), it was used in order to preserve the same killing method in all the
campaigns in this long-term experiment under conservation tillage management.
All of the CC residues were left on top of the soil. Thereafter, a new set of
N fertilizer treatments was set up for the maize cash crop phase. Maize
(Zea mays L., Pioneer P1574, FAO Class 700) was directly drilled on
7 April 2014 in all plots, resulting in a plant population density of
7.5 plants m-2; harvesting took place on 25 September
2014. The fertilizer treatments consisted of ammonium nitrate applied on
2 June at three rates: 170, 140 and 190 kg N ha-1 in F, V and B
plots, respectively, according to ISFM practices. For the calculation of each
N rate, the N available in the soil (which was calculated following soil
analysis as described below), the expected N uptake by maize crop, and the
estimated N mineralized from V and B residues were taken into account,
assuming that crop requirements were 236.3 kg N ha-1 (Quemada et al.,
2014). Estimated N use efficiency of maize plants for calculating N
application rate was 70 % according to the N use efficiency obtained
during the previous years in the same experimental area. Each plot received P
as triple superphosphate (45 % P2O5,
Fertiberia®, Madrid, Spain) at a rate of
69 kg P2O5 ha-1, and K as potassium chloride (60 %
K2O, Fertiberia®, Madrid, Spain) at a
rate of 120 kg K2O ha-1 just before sowing maize. All N, P and K
fertilizers were broadcast by hand, and immediately after N fertilization the
field was irrigated to prevent ammonia volatilization. The main crop previous
to sowing CCs was sunflower (Helianthus annuus L. `Sambro'). Neither the
sunflower nor the CCs were fertilized.
In order to determine the amount of N2O derived from the N fertilizers,
double-labeled ammonium nitrate (15NH415NO3, 5 at.%
15N, from Cambridge Isotope Laboratories, Inc., Massachusetts,
USA) was applied on 2 m × 2 m subplots established within each plot at a rate
of 130 kg N ha-1. In order to reduce biases due to the use of
different N rates (e.g., apparent priming effects or different mixing ratios
between the added and resident soil N pools) the same amount of N was
applied for all treatments. In each subplot, the CC residue was also left on
top of the soil. This application took place on 26 May by spreading
the fertilizer homogenously with a hand sprayer, followed by an irrigation
event.
Sprinkler irrigation was applied to the maize crop at a total amount of
688.5 mm in 31 irrigation events. Sprinklers were installed in a 12 m × 12 m
framework. The water doses to be applied were estimated from the crop
evapotranspiration (ETc) of the previous week (net water requirements). This
was calculated daily as ETc = Kc × ETo, where ETo is reference
evapotranspiration calculated by the FAO Penman–Monteith method (Allen et
al., 1998) using data from the meteorological station located in
the experimental field. The crop coefficient (Kc) was obtained using the
relationship for maize in semiarid conditions (Martínez-Cob, 2008).
Two different periods were considered for data reporting and analysis:
Period I (from CC sowing to N fertilization of the maize crop) and Period
II (from N fertilization of maize to the end of the experimental period,
after maize harvest).
GHG emissions sampling and analyzing
Fluxes of N2O, CH4 and CO2 were measured from October 2013 to
October 2014 using opaque, manually operated circular static chambers as described in
detail by Abalos et al. (2013). One chamber (diameter 35.6 cm,
height 19.3 cm) was located in each experimental plot. The chambers were
hermetically closed (for 1 h) by fitting them into stainless steel rings,
which were inserted at the beginning of the study into the soil to a depth
of 5 cm in order to minimize the lateral diffusion of gases and to avoid the soil
disturbance associated with the insertion of the chambers in the soil. The
rings were only removed during management events. Each chamber had rubber
sealing tape to guarantee an airtight seal between the chamber and the ring
and was covered with a radiant barrier reflective foil to reduce temperature
gradients between inside and outside. A rubber stopper with a three-way stopcock
was placed in the wall of each chamber to take gas samples. Greenhouse gas
measurements were always made with barley/vetch plants inside the chamber.
During the maize period, gas chambers were set up between maize rows.
During Period I, GHGs were sampled weekly or every 2 weeks. During the first
month after maize fertilization, gas samples were taken twice per week.
Afterwards, gas sampling was performed weekly or fortnightly, until the end
of the cropping period. To minimize any effects of diurnal variation in
emissions, samples were always taken at the same time of day
(10:00–12:00), which is reported as a representative time (Reeves and Wang, 2015).
Measurements of N2O, CO2 and CH4 emissions were made at 0, 30
and 60 min to test the linearity of gas accumulation in each chamber. Gas
samples (100 mL) were removed from the headspace of each chamber by syringe
and transferred to 20 mL gas vials sealed with a gastight neoprene septum.
The vials were previously flushed in the field using 80 mL of the gas
sample. Samples were analyzed by gas chromatography using a HP-6890 gas
chromatograph equipped with a headspace autoanalyzer (HT3), both from
Agilent Technologies (Barcelona, Spain). Inert gases were separated by HP
Plot-Q capillary columns. The gas chromatograph was equipped with a
63Ni electron-capture detector (micro-ECD) to analyze N2O
concentrations, and with a flame ionization detector (FID) connected to a
methanizer to measure CH4 and CO2 (previously reduced to
CH4). The temperatures of the injector, oven and ECD were 50, 50 and
350 ∘C, respectively. The accuracy of the gas
chromatographic data was 1 % or better. Two gas standards comprising a
mixture of gases (high standard with 1500 ± 7.50 ppm CO2,
10 ± 0.25 ppm CH4 and 2 ± 0.05 ppm N2O and low standard
with 200 ± 1.00 ppm CO2, 2 ± 0.10 ppm CH4 and 200 ± 6.00 ppb N2O)
were provided by Carburos Metálicos S.A. and
Air Products SA/NV, respectively, and used to determine a standard curve for
each gas. The response of the GC was linear within 200–1500 ppm for
CO2 and 2–10 ppm CH4 and quadratic within 200–2000 ppb for
N2O.
The increases in N2O, CH4 and CO2 concentrations within the
chamber headspace were generally (80 % of cases) linear
(R2 > 0.90) during the sampling period (1 h). Therefore,
emission rates of fluxes were estimated as the slope of the linear
regression between concentration and time (after corrections for
temperature) and from the ratio between chamber volume and soil surface area
(MacKenzie et al., 1998). Cumulative N2O, CH4 and CO2
emissions per plot during the sampling period were estimated by linear
interpolations between sampling dates, multiplying the mean flux of two
successive determinations by the length of the period between sampling and
adding that amount to the previous cumulative total (Sanz-Cobena et al.,
2014). The measurement of CO2 emissions from soil, including plants in
opaque chambers, only includes ecosystem respiration and not photosynthesis
(Meijide et al., 2010).
15N isotope analysis
Gas samples from the subplots receiving double-labeled AN fertilizer were
taken after 60 min of static chamber closure 1, 4, 9, 11, 15, 18, 22 and 25 days
after fertilizer application. Stable 15N isotope analysis of
N2O contained in the gas samples was carried out on a cryo-focusing gas
chromatography unit coupled to a 20/20 isotope ratio mass spectrometer (both
from SerCon Ltd., Crewe, UK). Ambient samples were taken occasionally as
required for the subsequent isotopic calculations. Solutions of 6.6 and 2.9 at. %
ammonium sulfate [(NH4)2SO4] were prepared and used
to generate 6.6 and 2.9 at. % N2O (Laughlin et al., 1997), which
were used as reference and quality control standards. In order to calculate
the atom percent excess (APE) of the N2O emitted in the subplots, the mean natural abundance
of atmospheric N2O from the ambient samples (0.369 at. % 15N)
was subtracted from the measured enriched gas samples. To obtain the N2O
flux that was derived from fertilizer (N2O-Ndff), the
following equation was used (Senbayram et al., 2009):
N2O-Ndff=N2O-N×N2O_APEsampleAPEfertilizer,
in which “N2O-N” is the N2O emission from soil,
“N2O_APEsample” is the 15N at. % excess of emitted
N2O, and “APEfertilizer” is the 15N at. % excess
of the applied fertilizer (Senbayram et al., 2009).
Soil and crop analyses
In order to relate gas emissions to soil properties, soil samples were
collected at 0–10 cm depth during the growing season on almost all
gas-sampling occasions, particularly after each fertilization event. Three
soil cores (2.5 cm diameter and 15 cm length) were randomly sampled close
to the ring in each plot, and then mixed and homogenized in the laboratory.
Soil NH4+ and NO3- concentrations were analyzed using 8 g of
soil extracted with 50 mL of KCl (1 M), and measured by automated
colorimetric determination using a flow injection analyzer (FIAS 400 Perkin
Elmer) provided with a UV-visible spectrophotometer detector. Soil (DOC) was
determined by extracting 8 g of homogeneously mixed soil with 50 mL of
deionized water (and subsequently filtered) and analyzed with a total organic
C analyzer (multi N/C 3100 Analityk Jena) equipped with an IR detector. The
water-filled pore space (WFPS) was calculated by dividing the volumetric
water content by total soil porosity. Total soil porosity was calculated
according to the following relationship: soil porosity = (1 - soil bulk
density/2.65), assuming a particle density of 2.65 g cm-3 (Danielson
and Sutherland, 1986). Gravimetric water content was determined by
oven-drying soil samples at 105 ∘C with a
Sartorius® MA30.
Four 0.5 m × 0.5 m squares were randomly harvested from each plot
before killing the CC by applying glyphosate. Aerial biomass was cut by hand
at soil level, dried, weighed and ground. A subsample was taken for
determination of total N content. From these samples the CC
biomass and N contribution to the subsequent maize were determined.
At maize harvest, two 8 m central rows in each plot were collected and
weighed in the field following separation of grain and straw. For
aboveground N uptake calculations, N content was determined in subsamples of
grain and biomass. Total N content of maize and CC subsamples was
determined with an elemental analyzer (TruMac CN, Leco).
Calculations and statistical analysis
Yield-scaled N2O emissions and N surplus in the maize cash crop were
calculated as the amount of N2O emitted (considering the emissions of
the whole experiment, i.e., Period I and Period II) per unit of aboveground N
uptake and taking the difference between N application and aboveground N
uptake, respectively (van Groenigen et al., 2010).
Statistical analyses were carried out with Statgraphics Plus 5.1. Analyses
of variance were performed for all variables during the experiment (except
climatic ones), for both periods indicated in Sect. 2.2. Data distribution
normality and variance uniformity were previously assessed by the Shapiro–Wilk
test and Levene's statistic, respectively, and transformed (log10,
root square, arcsin or inverse) before analysis when necessary. Means of
soil parameters were separated by Tukey's honest significance test at
P < 0.05, while cumulative GHG emissions, yield-scaled
N2O emissions and N surplus were compared by the orthogonal contrasts
method at P < 0.05. For non-normally distributed data, the
Kruskal–Wallis test was used on non-transformed data to evaluate
differences at P < 0.05. Linear correlations were carried
out to determine relationships between gas fluxes and WFPS, soil
temperature, DOC, NH4+ and NO3-. These analyses were
performed using the mean/cumulative data of the replicates of the CC
treatments (n= 12), and also for all the dates when soil and GHG were
sampled, for Period I (n= 16), Period II (n= 11) and the whole
experimental period (n= 27).
Discussion
Role of CCs in N2O emissions: Period I
Cover crop treatments (V and B) increased N2O losses compared to F,
especially in the case of V (Table 1). These results are consistent with the
meta-analysis of Basche et al. (2014), which showed that, overall,
CCs increase N2O fluxes (compared to bare fallow), with highly
significant increments in the case of legumes and a lower effect in the case
of nonlegume CCs. In the same experimental area, Sanz-Cobena et al. (2014)
found that V was the only CC significantly affecting N2O emissions. The
greatest differences between treatments were observed at the beginning
(13–40 days after CC sowing) and at the end of this period (229 days after
CC sowing) (Fig. 3a). On these dates, the mild soil temperatures and the
relatively high moisture content were more suitable for soil biochemical
processes, which may trigger N2O emissions (Fig. 1a, b) (Firestone and
Davidson, 1989). Average topsoil NO3- was significantly higher in
V (Fig. 2b), which was the treatment that led to the highest N2O
emissions. Legumes such as V are capable of biologically fixing atmospheric
N2, thereby increasing soil NO3- content with the
potential to be denitrified. Furthermore, the mineralization of the most
recalcitrant fraction of the previous V residue (which supplies nearly 4
times more N than the B residue, as indicated in Sect. 3.1.2) together
with high C-content sunflower residue could also explain higher
NO3- contents in V plots (Frimpong et al., 2011) and higher
N2O losses from denitrification (Baggs et al., 2000). After the CC kill
date, N release from decomposition of roots and nodules and faster
mineralization of V residue compared to that of B (shown by NO3-
in soil in Fig. 2c) are the most plausible explanations for the N2O
increases at the end of the intercrop period (Fig. 3a) (Rochette and Janzen,
2005; Wichern et al., 2008).
Some studies (e.g., Justes et al., 1999; Nemecek et al., 2008) have
pointed out that N2O losses can be reduced with the use of CCs, due to
the extraction of plant-available N unused by previous cash crop. However,
in our study lower N2O emissions were measured from F plots without CCs
during the intercrop period. This may be a consequence of higher
NO3- leaching in F plots (Gabriel et al., 2012; Quemada et
al., 2013), limiting the availability of the substrate for
denitrification. Frequent rainfall during the intercrop period (Fig. 1a) and
the absence of N uptake by CCs may have led to N losses through leaching,
resulting in low concentrations of soil mineral N in F plots.
Nitrous oxide emissions were low during this period but in the range of those
reported by Sanz-Cobena et al. (2014) in the same experimental area. Total
emissions during Period I represented 8, 10 and 21 % of total cumulative
emissions in F, B and V, respectively (Table 1). The absence of N fertilizer
application to the soil combined with the low soil temperatures during winter
– which were far from the optimum values for nitrification and
denitrification (25–30 ∘C) processes (Ussiri and Lal, 2013) – may
have caused these low N2O fluxes. The significant positive correlation
between soil temperature and N2O fluxes during this period highlights
the key role of this parameter as a driver of soil emissions (Schindlbacher
et al., 2004; García-Marco et al., 2014).
Role of CCs in N2O emissions: Period II
Isotopic analysis during Period II, in which ISFM was carried out, showed
that a significant proportion of N2O emissions came from endogenous
soil N or the mineralization of crop residues, especially after the first few
days following N fertilization (Fig. 4). In this sense, even though an
interaction between crop residue and N fertilizer application has been
previously described (e.g., in Abalos et al., 2013), the similar proportion
of N2O losses coming from fertilizer in B and F (without residue) 1
day after N fertilization revealed the importance of soil mineral N
contained in the micropores for the N2O bursts after the first
irrigation events, with respect to the N released from CC residues.
As we hypothesized, the different CCs played a key role in the N2O
emissions during Period II. Barley plots had higher N2O emissions than
fallow or V-residue plots (at the 10 % significance level; Table 1).
Further, a higher proportion of N2O emissions was derived from the
fertilizer in B-residue than in V-residue plots (Fig. 4). These results are
in agreement with those of Baggs et al. (2003), who reported a higher
percentage of N2O derived from the 15N-labeled fertilizer using a
cereal (ryegrass) as surface mulching instead of a legume (bean), in a field
trial with zero-tillage management. The differences between B and V in terms
of cumulative N2O emissions and in the relative contribution of each
source to these emissions (fertilizer- or soil-N) could be explained by:
(i) the higher C : N residue of B (20.7 ± 0.7 while that of V was
11.1 ± 0.1, according to Alonso-Ayuso et al., 2014) may have provided
an energy source for denitrification (Sarkodie-Addo et al., 2003), favoring
the reduction of the NO3- supplied by the synthetic fertilizer and
enhancing N2O emissions, as supported by the positive correlation of DOC
with the proportion of N2O coming from the synthetic fertilizer;
(ii) NO3- concentrations, which tended to be higher in B during the
maize cropping phase, could have led to incomplete denitrification and larger
N2O / N2 ratios (Yamulki and Jarvis, 2002); (iii) the easily
mineralizable V residue (with low C : N ratio) provided an additional N
source for soil microorganisms, thus decreasing the relative amount of
N2O derived from the synthetic fertilizer (Baggs et al., 2000; Shan and
Yan, 2013); and (iv) V plots were fertilized with a lower amount of
immediately available N (i.e., ammonium nitrate) than B plots, which could
have resulted in better synchronization between N release and crop needs
(Ussiri and Lal, 2013) in V plots. Supporting these findings, Bayer et
al. (2015) recently concluded that partially supplying the maize N
requirements with winter legume cover crops can be considered a N2O
mitigation strategy in subtropical agro-ecosystems.
Proportion of N2O losses (%) that come from N synthetic
fertilizer during Period II, for the three CC treatments (fallow, F;
vetch, V; and barley, B). Vertical lines indicate standard errors. “NS” and
* denote not significant and significant at P < 0.05,
respectively.
The mineralization of B residues resulted in higher DOC contents for these
plots compared to the F or V plots (P < 0.001). This was observed
in both Period I (as a consequence of soil C changes after the 8-year
cover-cropping management) and Period II (due to the CC decomposition).
Although in the present study the correlation between DOC and N2O
emissions was not significant, positive correlations have been previously
found in other low-C Mediterranean soils (e.g., Vallejo et al., 2006;
López-Fernández et al., 2007). Some authors have suggested that
residues with a high C : N ratio can induce microbial N immobilization
(Frimpong and Baggs, 2010; Dendooven et al., 2012). In our experiment, a
N2O peak was observed in B plots 20–25 days after fertilization
(Fig. 3b) after a remarkable increase of NO3- content (Fig. 2d),
which may be a result of a remineralization of previously immobilized N in
these plots.
The positive correlation of N2O fluxes and soil NO3-
content and WFPS during the whole cycle further supports the importance of
denitrification process for explaining N2O losses in this
agro-ecosystem (Davidson et al., 1991; García-Marco et al., 2014).
However, the strong positive correlation of N2O with NH4+
indicated that nitrification was also a major process leading to N2O
fluxes, and showed that the continuous drying-wetting cycles during a summer
irrigated maize crop in a semiarid region can lead to favorable WFPS
conditions for both nitrification and denitrification processes (Fig. 1c)
(Bateman and Baggs, 2005). Emission factors ranged from 0.2 to 0.6 % of
the synthetic N applied, which were lower than the IPCC default value of
1 %. As explained above, ecological conditions during the intercrop period
(rainfall and temperature) and maize phase (temperature) could be considered
normal (based on the 30-year average) in Mediterranean areas. Aguilera et
al. (2013) obtained a higher emission factor for high (1.01 %)
and low (0.66 %) water-irrigation conditions in a meta-analysis of
Mediterranean cropping systems. We hypothesized that management practices
may have contributed to these low emissions, but other inherent factors such
as soil pH should also be considered. Indeed, a higher N2O / N2
ratio has been associated with acidic soils, so lower N2O emissions from
denitrification could be expected in alkaline soils (Mørkved et al.,
2007; Baggs et al., 2010).
Methane and CO2 emissions
As is generally found in non-flooded arable soils, all treatments were net
CH4 sinks (Snyder et al., 2009). No significant differences were
observed between treatments in any of the two periods (Table 1), which is
similar to the pattern observed by Sanz-Cobena et al. (2014). Some
authors (Dunfield and Knowles, 1995; Tate, 2015) have suggested an
inhibitory effect of soil NH4+ on CH4 uptake. Low
NH4+ contents during almost all of the CC and maize cycle may
explain the apparent lack of this inhibitory effect (Banger et al., 2012).
However, during the dates when the highest NH4+ contents were
reached in V and B (225 days after CC sowing) (Fig. 3a), CH4 emissions
were significantly higher for these plots (0.12 and 0.16 mg CH4-C m-2 d-1
for V and B, respectively) than for F (-0.01 mg CH4-C m-2 d-1)
(data not shown). Similarly, the NH4+ peak
observed 2 days after fertilization (Fig. 3b) decreased in the order
V > F > B, the same trend as CH4 emissions
(which were 0.03, -0.04 and -0.63 mg CH4-C m-2 d-1 in
V, F and B, respectively; data not shown). Contrary to Sanz-Cobena et al. (2014), the presence of CCs did not
increase CO2 fluxes (Table 1) during the whole of Period I (which was
longer than the period considered by these authors), even though higher
fluxes were associated with B (but not V) with respect to F plots in the last
phase of the intercrop. This was probably as a consequence of higher root
biomass and plant respiration rates in the cereal (B) than in the legume (V).
Differences from fall to early winter were not significant, since low soil
temperatures limited respiration activity. The decomposition of CC residues
and the growth of the maize rooting system resulted in an increase in
CO2 fluxes during Period II (Oorts et al., 2007; Chirinda et al., 2010),
although differences between treatments were not observed.
Yield-scaled emissions, N surplus and general assessment
Yield-scaled N2O emissions ranged from 1.74 to 7.15 g N2O-N kg
aboveground N uptake-1, which is about 1–4 times lower than those
reported in the meta-analysis of van Groenigen et al. (2010) for a
fertilizer N application rate of 150–200 kg ha-1. Mean N surpluses of V
and F (Table 1) were in the range (0–50 kg N ha-1) recommended by van
Groenigen et al. (2010), while the mean N surplus in B (55 kg N ha-1)
was also close to optimal. In spite of higher N2O emissions
in V during Period I (which accounted for a low proportion of total
cumulative N2O losses during the experiment), these plots did not emit
greater amounts of N2O per kg of N taken up by the maize plants, and
even tended to decrease yield-scaled N2O emissions and N surplus (Table 1).
Adjusting fertilizer N rate to soil endogenous N led to lower N2O
fluxes than previous experiments where conventional N rates were applied
(e.g., Adviento-Borbe et al., 2007; Hoben et al., 2011; Sanz-Cobena et al.,
2012; Li et al., 2015), in agreement with the study by Migliorati et al. (2014).
Moreover, CO2 equivalent emissions associated with
manufacturing and transport of N synthetic fertilizers (Lal, 2004) can be
reduced when low synthetic N input strategies, such as ISMF, are employed.
Our results highlight the critical importance of the cash crop period on
total N2O emissions and demonstrate that the use of nonlegume
and – particularly – legume CCs combined with ISFM may provide an optimum
balance between GHG emissions from crop production and agronomic efficiency
(i.e., lowering synthetic N requirements for a subsequent cash crop, and
leading to similar yield-scaled N2O emissions as fallow).
The use of CCs has environmental implications beyond effects on direct soil
N2O emissions. For instance, CCs can mitigate indirect N2O losses
(from NO3- leaching). In the study by Gabriel et al. (2012),
conducted in the same experimental area, NO3- leaching was
reduced (on average) by 30 and 59 % in V and B, respectively.
Considering an emission factor of 0.075 from N leached (De Klein et
al., 2006), indirect N2O losses from leaching could be
mitigated by 0.23 ± 0.16 and 0.45 ± 0.17 kg N ha-1 yr-1
if V and B are used as CCs, respectively. Furthermore, the recent
meta-analysis of Poeplau and Don (2015) revealed a C sequestration potential
of 0.32 ± 0.08 Mg C ha-1 yr-1 with the introduction of CCs.
These environmental factors, together with CO2 emissions associated with
CC sowing and killing, should be assessed in future studies in order to
confirm the potential in CCs for increasing both the agronomic and the
environmental efficiency of irrigated cropping areas.