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Global Change Biology
Global Change Biology (2015) 21, 1249–1257, doi: 10.1111/gcb.12802
How inhibiting nitrification affects nitrogen cycle and
reduces environmental impacts of anthropogenic nitrogen
input
CHUNLIAN QIAO1,2, LINGLI LIU1, SHUIJIN HU3, JANA E. COMPTON4,
T A R A L . G R E A V E R 5 and Q U A N L I N L I 6
1
State Key Laboratory of Vegetation and Environmental Change, Institute of Botany, Chinese Academy of Sciences, No. 20
Nanxincun, Xiangshan, Beijing 100093, China, 2University of Chinese Academy of Sciences, No. 19 Yuquan Road, Beijing
100049, China, 3Department of Plant Pathology, North Carolina State University, Raleigh, NC 27695, USA, 4Western Ecology
Division, US Environmental Protection Agency, Corvallis, OR 97333, USA, 5National Center for Environmental Assessment, US
Environmental Protection Agency, Research Triangle Park, NC 27695, USA, 6Biostatistics and Bioinformatics, Department of
Medicine, Cedars-Sinai Medical Center, Los Angeles, CA 90048, USA
Abstract
Anthropogenic activities, and in particular the use of synthetic nitrogen (N) fertilizer, have doubled global annual
reactive N inputs in the past 50–100 years, causing deleterious effects on the environment through increased N leaching and nitrous oxide (N2O) and ammonia (NH3) emissions. Leaching and gaseous losses of N are greatly controlled
by the net rate of microbial nitrification. Extensive experiments have been conducted to develop ways to inhibit this
process through use of nitrification inhibitors (NI) in combination with fertilizers. Yet, no study has comprehensively
assessed how inhibiting nitrification affects both hydrologic and gaseous losses of N and plant nitrogen use efficiency. We synthesized the results of 62 NI field studies and evaluated how NI application altered N cycle and ecosystem services in N-enriched systems. Our results showed that inhibiting nitrification by NI application increased
NH3 emission (mean: 20%, 95% confidential interval: 33–67%), but reduced dissolved inorganic N leaching (48%,
56% to 38%), N2O emission (44%, 48% to 39%) and NO emission (24%, 38% to 8%). This amounted to a
net reduction of 16.5% in the total N release to the environment. Inhibiting nitrification also increased plant N recovery (58%, 34–93%) and productivity of grain (9%, 6–13%), straw (15%, 12–18%), vegetable (5%, 0–10%) and pasture
hay (14%, 8–20%). The cost and benefit analysis showed that the economic benefit of reducing N’s environmental
impacts offsets the cost of NI application. Applying NI along with N fertilizer could bring additional revenues of
$163 ha1 yr1 for a maize farm, equivalent to 8.95% increase in revenues. Our findings showed that NIs could create
a win-win scenario that reduces the negative impact of N leaching and greenhouse gas production, while increases
the agricultural output. However, NI’s potential negative impacts, such as increase in NH3 emission and the risk of
NI contamination, should be fully considered before large-scale application.
Keywords: cost–benefit analysis, ecosystem services, N2O emission, NH3 emission, nitrogen fertilizer, nitrogen leaching,
nitrogen management, NO emission
Received 5 May 2014; revised version received 21 October 2014 and accepted 21 October 2014
Introduction
Synthetic nitrogen fertilizers were developed via the
Haber–Bosch process over a century ago and have been
applied intensively since then to increase plant growth
(Galloway et al., 2008); they now exceed terrestrial biological N fixation as the main source of new reactive N
(Nr) to the global N cycle (Erisman et al., 2011; Fowler
et al., 2013; Vitousek et al., 2013). N fertilizers have
made a remarkable contribution in the alleviation of
global food shortage, increasing food production by
almost 50% (Sutton et al., 2011). However, N fertilizers
Correspondence: Lingli Liu, tel. +86 10 62836160,
fax +86 10 82596146, e-mail: [email protected]
© 2014 John Wiley & Sons Ltd
not taken up by the target system tend to mobilize,
causing serious environmental consequences as the N
cascades into the atmosphere, terrestrial and aquatic
systems (Sutton et al., 2011; De Vries et al., 2013; Fowler
et al., 2013).
In most regions, about 20–70% of the fertilizer is lost
to the environment as dissolved N, greenhouse gases
(GHG), for example, N2O, or other air pollutants, for
example, NH3 and NO (Smil, 1999; Bouwman et al.,
2009; Sutton et al., 2011). The massive release of the
excess Nr greatly disturbs the natural biogeochemical
cycle of N, resulting in severe environmental problems
in water, air and soil (Davidson et al., 2012; De Vries
et al., 2013; Fowler et al., 2013). Taking China as an
1249
1250 C . Q I A O et al.
example, N fertilizer contributed 66% of the country’s
N2O emission, caused severe eutrophication of 62% of
total freshwater areas (Ju et al., 2004; Le et al., 2010) and
led to atmospheric N deposition as high as 89–
104 kg ha1 yr1 in intensive agricultural regions (Ju
et al., 2009). Similarly, in Europe, anthropogenic nitrogen contributed to 10% of the global N2O emission and
80% of European fresh waters exceed the level of
1.5 mg N L1, which is the threshold for high risk of
biodiversity loss (Sutton et al., 2011). Taken together,
the environmental problems caused by anthropogenic
N threaten global sustainability (Rockstr€
om et al., 2009;
De Vries et al., 2013; Fowler et al., 2013). Maintaining
food production while reducing the detrimental effects
of anthropogenic N application is an urgent priority for
global food security and environmental sustainability
(Giller et al., 2004; Erisman et al., 2011; Davidson et al.,
2012).
When N fertilizers are applied, usually as urea or
anhydrous ammonia, a microbial process called nitrification converted most N fertilizers into the highly
mobile NO
3 within 2–3 weeks (Huber et al., 1977),
which causes low retention in the target system. Most
NO
3 losses occur through leaching or denitrification
before plants can utilize it; thus, the system has low
nitrogen use efficiency (NUE). Nitrification inhibitors
(NIs) are a group of chemical compounds that suppress
the first step of nitrification by inhibiting Nitrosomonas
spp. bacteria that oxidize NHþ
4 to nitrite (NO2 ), and
therefore delay the nitrification process (Huber et al.,
1977; Zerulla et al., 2001). Inhibiting nitrification
through use of NI can encourage the retention of soil N
as the less mobile NHþ
4 form, which could significantly
reduce leaching loss of NO
3 from soil. Lower soil NO3
concentration resulting from the inhibition of nitrification also reduces substrate availability for denitrification and decreases N2O emissions (Akiyama et al.,
2010). NIs are recommended by IPCC as a potential
mitigation option for cutting agricultural N2O emissions (IPCC, 2014). However, inappropriate use of NI
can cause adverse impacts. NIs are often applied to
pastures to reduce GHG emission and N leaching. NIs
degrade over time, but can be a concern if livestock are
fed with hay containing NI residues, as happened in
New Zealand in 2012. Trace amounts of dicyandiamide
(DCD) were found in some dairy products, raising
great concerns about the safe use of NI in feedlot areas
(MPI, 2013).
Hundreds of field and laboratory studies have investigated how inhibiting nitrification through use of NIs
affected N cycling and plant productivity since the
1960s. Although the effects varied depending on factors
such as soil texture, physicochemical characters of NIs
and plant species, appropriate applications of NI with
N fertilizer often increased plant NUE and alleviate
environmental damages caused by N fertilizers (Huber
et al., 1977; Giller et al., 2004). To better understand
how inhibiting nitrification by NI application affects
the environmental impacts caused by N fertilizers, we
conducted a comprehensive meta-analysis including
soil acidification, N leaching, air pollutant emission,
GHG emission and plant productivity. We also
assessed whether the responses of those variables will
be altered by NI forms, ecosystem types, fertilizer types
and soil texture. In addition, we performed a cost–benefit analysis (CBA) based on ecosystem services following studies by Compton et al. (2011) and Van Grinsven
et al. (2013) and quantified the overall impacts of
NI application on ecosystem services in agricultural
systems by monetary values.
Materials and methods
Data selection
Peer-reviewed NI studies were searched using Web of Science (1984–2013). The impacts of NI were assessed by the
following six categories with 19 variables, including (i) soil
chemistry: soil pH, and the concentrations of extractable
NHþ
4 , extractable NO3 and dissolved inorganic N (DIN);
(ii) N leaching: the leaching of NHþ
4 , NO3 and DIN; (iii)
GHG emission: the emissions of N2O and CH4, and the
uptake of CH4; (iv) air pollutant emission: the emissions
of NO and NH3; (v) crop productivity: the productivity of
grain, straw, vegetable and pasture hay; and (vi) NUE
of plants: N concentration in plant tissue, plant N uptake,
which was defined as the total N content in the harvested
aboveground biomass per unit area, and N recovery (RN),
which was defined as RN = (PN-P N0)/N, where PN was
plant N uptake (kg N/ha) in N fertilized plots , PN0 was
plant N uptake (kg N/ha) in control plots, N was the N
fertilization rate (kg N/ha).
The control and treatment means of the 19 assessed variables, including NI forms [DCD, 3, 4-Dimethylpyrazole phosphate (DMPP), nitrapyrin, Ca-carbide and organic NI],
ecosystem types (agriculture and pasture), fertilizer types
(synthetic N fertilizer, manure and the mixture of the both)
and soil texture classes (sand, loam, clay and silt), were
extracted from the available studies. Where data were presented graphically, figures were digitized using software ENGAUGE DIGITIZER (Free Software Foundation, Inc., Boston, MA,
USA). In addition, experimental location, latitude, plant species, NI and N manipulation level were also included in the
dataset. Details of the dataset structure were described in
Table S1 (See Supporting information).
Meta-analysis
In total, 62 peer-reviewed publications with 859 datasets
across the world were selected for our analysis (Fig. S1 and
© 2014 John Wiley & Sons Ltd, Global Change Biology, 21, 1249–1257
N I T R I F I C A T I O N I N H I B I T O R S F O R N I T R O G E N M A N A G E M E N T 1251
Table S1). For parameters of soil chemistry, including soil pH
and concentrations of NHþ
4 , NO3 and DIN, we used the average value across a whole crop rotation. For variables showing
great seasonal variability, such as the leaching loss of NHþ
4,
NO
3 and DIN, or the emissions of N2O, CH4, NO and NH3,
we only included studies which reported the cumulative values for a whole crop rotation, or the cumulative values for a
whole crop rotation can be estimated from the reported data.
If data of multiple crop rotations with same vegetation and NI
treatments were reported in a study, data of the latest crop
rotation were chosen because the meta-analysis model
required independence between observations. Measurements
were considered as independent observations if different NI
forms, fertilizer types or addition levels were manipulated in
the same experiment.
The response ratio (RR) was used to measure the effect of NI
application on the assessed variables (Hedges et al., 1999). For
each variable, the natural log of RR is calculated by: ln RR = ln
NþNI =X
N ), where X
N is the mean of the variable when N fer(X
NþNI is the mean of the variable
tilizer was applied alone, and X
when N fertilizer was applied with NI. Among 859 field data
entries in our dataset, only 266 reported standard errors/deviations of the experimental and control group responses.
Adams et al. (1997) developed resampling methods for metaanalysis of ecological data. Different from conventional parametric methods, resampling methods do not require computation of standard errors of the individual effect size estimations.
Studies that did not report standard errors/deviations therefore can be included in the analysis (Adams et al., 1997). We
adopted the resampling methods and estimated the mean RR
with bias-corrected 95% confidential intervals (CI) using bootstrap resampling procedure (Adams et al., 1997).
For each variable, total heterogeneity among the categorical
group (QT) was partitioned into within-group heterogeneity
(QW) and between-group heterogeneity (QB), where QT, QW
and QB have chi-square distributions. A significance of QB
indicates that the RRs are significantly different between the
levels of the categorical group. Publication bias was assessed
with studies that reported standard errors/deviations. We
used Egger’s regression test to quantify publication bias,
which has higher power to detect the bias than the rank correlation method (Sterne et al., 2001).
Meta-analysis was performed by the software METAWIN 2.1
(Sinauer Associates, Inc. Sunderland, MA, USA). The response
of each variable was considered significant if the 95% CI of RR
did not overlap 1.
of FN among experimental conditions such as climate, soil
type, cropping time, seasonality fertilization and fertilization
method.
To evaluate whether NI application is a feasible approach
to reduce N’s environmental impacts and increase agricultural
outputs, a CBA was conducted using a US maize farm as a
case study. The net economic impact of NI was assessed by
summing the cost/benefit associated with (i) NI’s impacts on
ecosystem services, including DIN leaching, GHG emission
and air pollutant emission; (ii) NI’s impacts on crop productivity; and (iii) the expense of purchasing NI. For each individual ecosystem service, NI’s impact on the economic value (M)
of this service is estimated by the formula: M = N 9
(FN+NI FN) 9 P. Here, N is a fertilizer input rate that is
125 kg N ha1 yr1, which is the mean annual N fertilizer
input rate in US cropland (FAO/IFA, 2001). FN+NI and FN are
defined above. P is the monetary value of the related environmental impacts given by previous CBA studies (Jaynes et al.,
2010; Kusiima & Powers, 2010; Compton et al., 2011), and the
details were listed in Table 2.
Results
Soil chemistry
Averaged across all studies, NI application raised soil
pH by 0.23 pH units (95% CI: 0.17–0.34, Table S2). NI
application significantly increased soil-extractable NHþ
4
concentration (41%, 27–60%) and decreased extractable
NO
3 concentration (41%, 48% to 33%, Fig. 1, Table
S2), although publication bias was detected for NO
3
concentration (Egger’s P-value = 0.05, Table S9). When
data were examined based on the chemical nature of
NIs, all NIs significantly increased NHþ
4 concentration,
except for nitrapyrin. All NIs significantly decreased
NO
3 concentration (Table S2).
Averaged across all studies, NI application did not
alter the concentration of DIN (Fig. 1, Table S2). However, when data were subdivided by NI types, DIN
decreased in soil applied with organic NI, but was not
affected by DCD, DMPP and other NIs (Table S2). DIN
increased (20%, 9–32%) in pasture soils, but decreased
(12%, 19% to 3%) in agricultural soils. Fertilizer
types and soil texture also altered DIN’s response to NI
addition (Table S2).
Nitrogen loss factor and cost–benefit analysis
N loss factor under conventional fertilizer practice (FN) is
defined as the ratio of the amount of N fertilizer lost to environment to the amount of N fertilizer applied to soil. FN values were derived from the literatures (FAO/IFA, 2001;
Bouwman et al., 2009; Liu et al., 2010). For the variable whose
RR was significantly altered by NI application, its N loss factor
under NI application (FN+NI) was estimated by multiplying FN
with the corresponding RR calculated from this meta-analysis.
Because of data limitation, we did not assess the heterogeneity
© 2014 John Wiley & Sons Ltd, Global Change Biology, 21, 1249–1257
N leaching
On average, NI reduced NO
3 leaching (47%, 59% to
32%) and total DIN leaching (48%, 56% to 38%)
(Fig. 1, Table S3). The reduction in NO
3 and DIN leaching was significant for different types of ecosystem, fertilizer and soil texture (Table S3). For NHþ
4 leaching,
the overall effect of NIs was not significant (Table S3).
However, when data were subdivided in different
1252 C . Q I A O et al.
responses among N fertilizer types were similar (Fig. 1,
Table S5). Averaged across all studies, NI increased
NH3 emission (20%, 7–33%, Fig. 1, Table S5), although
publication bias was detected (Egger’s P-value < 0.01,
Table S9). The responses of NH3 emission to NI addition varied for different NI forms, ecosystem types, fertilizer types and soil types. DCD increased NH3
emission (34%, 20–50%), but DMPP and Ca-carbide had
no significant impact on it (Table S5). NH3 emission
was significantly increased in agricultural soils (21%,
4–44%), but not in pasture soils. Fertilizer type and soil
texture also affected the responses of NH3, emission to
NI addition (Table S5).
Crop productivity
Fig. 1 Effects of nitrification inhibitors on soil chemistry, environment impacts, plant productivity and plant nitrogen use
efficiency.
ecosystem type, ammonium leaching increased in agricultural lands (25%, 17–34%), but decreased in pastures
(33%, 54% to 4%). NI’s impacts on NHþ
4 leaching
also differed among fertilizer types and soil texture
(Table S3).
Greenhouse gas emission from soil
When averaged across all studies, NI application significantly decreased N2O emission (44%, 48% to
39%). The reduction in N2O emission induced by NI
was significant for all types of ecosystem, fertilizer and
soil texture (Table S4). NI had no significant impact on
the emission and uptake of CH4 (Fig. 1, Table S4), and
the responses of CH4 emission and uptake were similar
under different experimental conditions, such as NI
form, ecosystem type, fertilizer type and soil texture
(Table S4).
Air pollutant emission from soil
Nitrification
inhibitors
application
significantly
decreased NO emission (24%, 38% to 8%), and the
Across all studies, NI applications significantly
increased productivity of grain (9%, 6–13%), straw
(15%, 12–18%), vegetables (5%, 0–10%) and pasture hay
(14%, 8–20%, Fig. 1, Table S6). Among the four most
commonly studied crops, grain productivity of barley
showed the greatest increase (17%, 7–29%), as compared to wheat (2%, 0–5%), maize (9%, 3–17%) and rice
(7%, 3–12%, Fig. 1, Table S6). For all crop systems, DCD
increased plant productivity except for vegetable.
DMPP increases plant productivity as well, except for
pasture hay (Table S6). In addition, the effects of NI on
crop production varied among different soil texture
classes (Table S6).
Nitrogen use efficiency
On average, NI increased plant N concentration (7%,
1–16%), N uptake (15%, 11–20%) and N recovery
(58%, 34–93%, Fig. 1, Table S7), although publication
bias was detected for N concentration (Egger’s
P-value < 0.01, Table S9). For different NI forms,
DMPP showed no significant effect on plant N concentration, and nitrapyrin showed no significant
impact on N recovery. When data were divided into
different ecosystem types, NI did not affect plant N
concentration in pastures. In general, the responses
of N concentration, N uptake and N recovery to NI
application were not affected by fertilizer types and
soil texture classes (Table S7).
Nitrogen loss factor and cost–benefit analysis
Compared to conventional N fertilizer practice, application of NI with N fertilizer increased N loss factor for
gaseous emission from 0.156 to 0.178, but decreased N
loss factor for DIN leaching from 0.154 to 0.081
(Table 1). Overall, NI application resulted in a decrease
in total N loss factor from 0.310 to 0.259 (Table 1),
© 2014 John Wiley & Sons Ltd, Global Change Biology, 21, 1249–1257
N I T R I F I C A T I O N I N H I B I T O R S F O R N I T R O G E N M A N A G E M E N T 1253
equivalent to a net reduction of the total N loss by
16.5% (Table 1).
The CBA case study indicated that NI application
brought a total environmental benefit of $25.16 ha1,
mostly by reducing the impacts associated with DIN
leaching and NO emission (Table 2). The increases
in maize yield increased the revenue by
$163.83 ha1 (Table 2). After deducting the cost of
purchasing NI, NI application resulted in a net
monetary benefit of $162.70 ha1 for a maize farm
(Table 2). Based on the average maize yield
(9.24 ton ha1) and its market price in 2012
($197 ton1), the mean revenue for a maize farm in
US is $1820 ha1. NI application will raise the revenue to $162.70 ha1, which is about 8.95% increase
if the environmental benefit was accounted.
Discussion
Previously, it has been unclear whether the benefits
of inhibiting nitrification through use of NI could
offset or even outweigh the cost because there had
not been a systematic analysis considering environmental effects, forms of NI and types of crop systems. Using meta-analysis and CBA, we conducted
a comprehensive assessment on how NI alters N
cycling and ecosystem services and evaluated
whether NI application is an effective approach for
improving N management.
Soil acidification
Soil acidification is one of the most common consequences of soil degradation caused by N overuse (Cui
et al., 2013; Zhang et al., 2013). Our results indicated
NI application increased soil pH by 0.23 units, suggesting that NI application could alleviate soil acidifi-
cation. The increase in soil pH under NI application
could be caused by the combination of the decelerated
rate of nitrification and increased plant N use efficiency, as indicated by higher N uptake (Fig. 1, Table
S7). Instead of loss through cation-depleting N leaching, most NO
3 released by nitrifiers was absorbed by
roots, which released OH and balanced soil pH. Considering that many farm fields add lime to optimize
pH for nutrient availability, the benefit of NI application could include a small reduction in lime additions
to the fields.
Nitrogen losses to environment
The greatest impacts of anthropogenic N addition
beyond the intended use are impairment of ecosystem services in adjacent and downstream environments related to soil, air and water pollution (De
Vries et al., 2013; Van Grinsven et al., 2013).
Although NI application did not change soil DIN
concentration, it altered DIN composition by increas
ing NHþ
4 concentration and reducing NO3 concentration in soil (Fig. 1). The change in inorganic N
composition altered the pathways of N losses to
environment and resulted in different impacts on
water and air. Higher soil NHþ
4 concentration led to
more NH3 volatilization, which may exaggerate NH3
pollution and result in more damage on human
health (Compton et al., 2011). On the other hand, NI
reduced the emission of NO and N2O by depressing
nitrification. However, the decrease in the amount of
N emitted as NO and N2O is less than the increase
in N emitted as NH3 (Table 1). Overall, NI application increased gaseous N loss factor by 14.3%
(Table 1).
Nitrate is highly mobile in soil and comprises the
major fraction of N leaching loss from agricultural sys-
Table 1 N loss factors with and without nitrification inhibitors (NI) application in croplands. RR: response ratio; FN: N loss factor
without NI; FN+NI: N loss factor with NI; Change in N loss: the percentage difference between N loss factor with and without NI.
Positive and negative values indicated the increase and decrease, respectively, in N loss when N fertilizer was applied with NI
Nitrogen loss
RR
FN
FN+NI
Change in N loss (%)
NH3 emission
N2O emission
NO emission
Dissolved inorganic N (DIN) leaching
Total gaseous N loss‡
Total N loss§
1.201
0.560
0.758
0.524
0.140*
0.010*
0.006*
0.154†
0.156
0.310
0.168
0.006
0.005
0.081
0.178
0.259
20.1
44.0
24.2
47.6
14.3
16.5
*The source of the data was FAO/IFA (2001).
†Calculated by using N budget data from three global N synthesis studies: Smil (1999), Bouwman et al. (2009) and Liu et al. (2010).
‡The sum of N loss through NH3, N2O and NO emission.
§The sum of N loss through NH3, N2O and NO emission and DIN leaching.
© 2014 John Wiley & Sons Ltd, Global Change Biology, 21, 1249–1257
1254 C . Q I A O et al.
Table 2 The cost/benefit analysis for a maize farm applying nitrification inhibitors (NI) with fertilizer rate of 125 kg N ha1 yr1.
For change in N loss under NI, positive values indicate that NI increases N losses, and negative ones indicate N reduces N loss. For
the monetary response, the positive numbers indicate the amount of the economic benefit, whereas the negative ones indicate the
amount of the economic cost
Variables
Assessed impacts
Cost (data source)
NH3 emission
The cost of human health
damage
The cost of climate change
$1.30 kg1 N (Compton
et al., 2011)
$1.24 kg1 N (Kusiima &
Powers, 2010)
$23.00 kg1 N (Compton
et al., 2011)
$2.71 kg1 N (Jaynes
et al., 2010)
N2O emission
NO emission
The cost human health
damage
Dissolved inorganic The abatement cost of
N leaching
reducing N from agricultural
drainage water
Sum of the environmental impacts
Change in N loss under NI
(kg N1 ha1)*
Monetary
response ($ha1)
3.52
4.58
0.55
0.69
0.18
4.18
9.17
24.84
25.16
Changes in yield
(ton ha1)#
Monetary
response ($ ha1)
Variables
Assessed impacts
Unit price (data source)
Maize production
The benefit of increase
in yield
$197.00 ton1
(USDA, 2013)
Variables
Assessed impacts
Unit price
Application rate
(kg ha1)†
Monetary
response ($ ha1)
$1.75 kg1¶
15.00
26.25
Dicyandiamide
The cost of
(DCD)
purchasing DCD
Sum of the monetary responses
0.83
163.83
162.70
*Changes in N loss under NI = 125 kg N ha1 9 (FN+NI FN). The positive value indicated an increase in N loss under NI, and
the negative values indicated a decrease in N loss under NI.
#The change in maize production was calculated by multiplying the mean maize production in US (9.24 ton ha1, USDA, 2012)
with mean response ratio of maize yield (1.09) estimated by the current study (Table S6).
†The recommended DCD application rate (15 kg ha1 yr1) was from Di & Cameron (2003).
¶The price of DCD was the mean of the market price of DCD from the website of Alibaba.
tems (Smil, 1999). The decrease in soil NO
3 concentration under NI addition resulted in a significant decrease
þ
in NO
3 leaching. Although soil NH4 concentration
increased under NI application, ammonium leaching
was not changed (Fig. 1), probably because NHþ
4 is not
as mobile as NO
3 . Compared to conventional N fertilizer practice, NI application reduced DIN loss factor by
47.6% (Table 1).
The environmental impacts of NI application
depend on whether NI could reduce the net amount
of N release to the environment. Our estimation
suggested that when NI was applied along with N
fertilizers, the decrease in N leaching outweighed the
increase in N gaseous loss. Overall, NI application
decreased total N loss by 16.5% (Table 1), suggesting
that NI application is an effective practice to reduce
N lost to environment. However, it should be aware
that many factors, such as climate, soil types, cropping
time, seasonality fertilization and fertilization method,
would affect the value of N loss factor (FN) and eventually alter the estimation for the amount of N loss.
Because of data limitation, we could not assess
how those factors will affect N loss factor under NI
application. More studies are needed to address the
uncertainties.
Greenhouse gas emissions
Agriculture accounts for 60% of N2O and 50% of
CH4 global anthropogenic emission (IPCC, 2007).
Synthetic N fertilizer is the largest contributor to
N2O emission in agricultural systems and may alter
emissions of CO2 and CH4 as well (Liu & Greaver,
2009). We found that NI application reduced N2O
emission by 44%, which is slightly more than the
38% reported in a previous meta-analysis (Akiyama
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N I T R I F I C A T I O N I N H I B I T O R S F O R N I T R O G E N M A N A G E M E N T 1255
et al., 2010). N2O is released during both nitrification
and denitrification processes. NI application not only
reduces N2O emission by inhibiting nitrification, it
also reduces emission via denitrification by reducing
soil NO
3 concentration. Our results also suggested
that NIs had no significant effect on soil CH4 emission and uptake. However, the small number of
field studies on CH4 emission and uptake limited
the statistical power of the analysis.
Plant productivity and NUE
Previous studies have not identified a consistent
effect of NI application on plant productivity. A
meta-evaluation on those studies is needed to assess
whether NI is an effective approach to improve crop
yield. Until now, only Wolt (2004) used meta-analysis to assess NI research trials in Midwest USA and
found that nitrapyrin increased maize production by
7%. In the current study, we assessed how NI
application affected productivity in different plant
systems.
Although the responses varied among crops, NI
application significantly increased the productivity of
grain, straw, vegetable and hay (Fig. 1, Table S6). The
increase in plant productivity was attributed to the
higher NUE under NI application, as indicated by
greater plant N uptake and N recovery rate (Fig. 1).
Higher NUE not only enhanced plant productivity, but
also partially reduced N’s environmental impact
(Table 1). With more N being recovered by plant, less
of the applied fertilizer was subjected to be lost to the
environment.
Nitrification inhibitors application caused a shift
in the relative proportion of NHþ
4 and NO3 in soil,
which may influence the plant response. Although
plants can take up both NHþ
4 and NO3 , species
may differ in their preference for the two chemical
forms of N (Li et al., 2013). Such preference could
affect species’ response to NI application. In the current study, we found that NI application induced
less increase in the productivity of vegetable, compared to that of grain, straw and hay. This is probably due to the fact that vegetables strongly prefer
þ
NO
3 to NH4 (Huber et al., 1977; Li et al., 2013). In
addition to enhancing NUE, NI application also
increased N concentration in plant tissue by 8%
(Fig. 1, Table S7). N concentration is an important
indicator for food and forage quality as N concentration is highly correlated with the concentration of
Fig. 2 Effects of nitrification inhibitors (NI) on plant productivity and other ecosystem services in N-amended ecosystems. Mean effects
of NI on each variable were shown in parentheses. NI, nitrification inhibitor; DIN, dissolved inorganic nitrogen.
© 2014 John Wiley & Sons Ltd, Global Change Biology, 21, 1249–1257
1256 C . Q I A O et al.
proteins. Higher N concentration in plant tissues
suggested that NI might further enhance the nutrition quality of crops.
As discussed above, NI could increase plant productivity and reduce N’s environmental impacts. Still, the
safety in use of NI should be assessed for different
plant systems before large-scale application. While
most studies focused on the application of NI in row
crops, there are concerns about application to pastures
and animal feed. The residence time of DCD on plant
canopy is between 6 and 16 days depending on
climate condition (Kim et al., 2012). Because dairy cows
ate grass before the sprayed DCD was completely
degraded, trace amounts of DCD were detected in
some milk products produced in New Zealand in
2012. Although DCD is of very low toxicity and the
concentration was lower than any level known to
cause human health concern, it raised great public
concerns and New Zealand Government suspended
the use of DCD (MPI, 2013). Appropriate application
procedures to avoid direct consumption of NI by
farm animals could greatly reduce the risk of
contamination.
local trials are needed to examine the effectiveness and
also the potential negative impacts of NI for particular
crops and in feedlots. Appropriate use of NI could help
reduce the cost of mitigating pollution and climate
change.
Implication for N policy development
Cui S, Shi Y, Groffman PM, Schlesinger WH, Zhu YG (2013) Centennial-scale analysis
of the creation and fate of reactive nitrogen in China (1910–2010). Proceedings of the
National Academy of Sciences of the United States of America, 110, 2052–2057.
Davidson E, David M, Galloway J et al. (2012) Excess nitrogen in the US environment:
trends, risks, and solutions. Issues in Ecology, 15, 1–16.
De Vries W, Kros J, Kroeze C, Seitzinger SP (2013) Assessing planetary and regional
Our results indicated that inhibiting nitrification by
NI application altered the pathways of the excess N
entering the environment, decreased overall loss and
increased the productivity of crops (Fig. 2).
Although the changes in the environmental pathways of N released from the system induced both
positive and negative environmental impacts, the
overall budget estimation suggests that NI application alleviates water pollution and lowers the global
warming potential induced by N fertilizer. Our CBA
suggests that the monetary benefit associated with
NI application outweighs the cost, resulting in an
increase in revenue of $162.70 ha1, which is equivalent to an increase of 8.95% in financial gain for a
maize farm.
Abatements of air pollution, water eutrophication
and global warming require significant investment of
public dollars. Our analysis suggests that inhibiting
nitrification by NI application is an effective approach
to increase agricultural output while also alleviating
the environmental damage caused by reactive N leaching and GHG production. NI did increase the amount
of N released to the environment as NH3, meaning that
NI application may exacerbate environmental costs in
areas with concerns about NH3 emissions. To encourage the use of NI, governments could design effective
policies, such as providing subsidies for farmers using
NIs. On the other hand, before large-scale application,
Acknowledgements
This study was supported financially by Chinese National Key
Development Program for Basic Research (2013CB956304,
2014CB954003) and National 1000 Young Talents Program. This
work has been subjected to EPA review and approved for publication. Mention of trade names or commercial products does
not constitute endorsement or recommendation for use.
References
Adams DC, Gurevitch J, Rosenberg MS (1997) Resampling tests for meta-analysis of
ecological data. Ecology, 78, 1277–1283.
Akiyama H, Yan X, Yagi K (2010) Evaluation of effectiveness of enhanced-efficiency
fertilizers as mitigation options for N2O and NO emissions from agricultural soils:
meta-analysis. Global Change Biology, 16, 1837–1846.
Bouwman A, Beusen A, Billen G (2009) Human alteration of the global nitrogen and
phosphorus soil balances for the period 1970–2050. Global Biogeochemical Cycles, 23,
GB0A04.
Compton JE, Harrison JA, Dennis RL et al. (2011) Ecosystem services altered by
human changes in the nitrogen cycle: a new perspective for US decision making.
Ecology Letters, 14, 804–815.
nitrogen boundaries related to food security and adverse environmental impacts.
Current Opinion in Environmental Sustainability, 5, 392–402.
Di HJ, Cameron KC (2003) Mitigation of nitrous oxide emissions in spray-irrigated
grazed grassland by treating the soil with dicyandiamide, a nitrification inhibitor.
Soil Use and Management, 19, 284–290.
Erisman JW, Galloway J, Seitzinger S, Bleeker A, Butterbach-Bahl K (2011) Reactive
nitrogen in the environment and its effect on climate change. Current Opinion in
Environmental Sustainability, 3, 281–290.
FAO/IFA (2001) Global Estimates of Gaseous Emissions of NH3, NO and N2O from Agricultural Land. Food and Agriculture Organization/International Fertilizer Industry
Association, Rome, Italy.
Fowler D, Coyle M, Skiba U et al. (2013) The global nitrogen cycle in the twenty-first
century. Philosophical Transactions of the Royal Society of London. Series B, Biological
Sciences, 368, 20130164.
Galloway JN, Townsend AR, Erisman JW et al. (2008) Transformation of the
nitrogen cycle: recent trends, questions, and potential solutions. Science, 320,
889–892.
Giller KE, Chalk P, Dobermann A et al. (2004) Emerging technologies to increase
the efficiency of use of fertilizer nitrogen. In: Agriculture and the Nitrogen
Cycle: Assessing the Impacts of Fertilizer Use on Food Production and the
ment (eds Mosier A, Delgado J, Cochran V, Valentine D, Parton
35–52. Island Press, Washington, DC, USA.
Hedges LV, Gurevitch J, Curtis PS (1999) The meta-analysis of response
experimental ecology. Ecology, 80, 1150–1156.
Huber D, Warren H, Nelson D, Tsai C (1977) Nitrification inhibitors: new
food production. BioScience, 27, 523–529.
EnvironW), pp.
ratios in
tools for
IPCC (2007) Climate Change 2007: Mitigation of Climate Change: Working Group III Contribution to the Fourth Assessment Report of the Intergovernmental Panel on Climate
Change. Cambridge University Press: Cambridge, UK and New York, NY, USA.
IPCC (2014) Climate Change 2014: Mitigation of Climate Change. Contribution of
Working Group III to the Fifth Assessment Report of the Intergovernmental Panel on
Climate Change. Cambridge University Press, Cambridge, UK and New York, NY,
USA.
© 2014 John Wiley & Sons Ltd, Global Change Biology, 21, 1249–1257
N I T R I F I C A T I O N I N H I B I T O R S F O R N I T R O G E N M A N A G E M E N T 1257
Jaynes D, Thorp K, James D (2010) Potential water quality impact of drainage water
management in the Midwest USA. American Society of Agricultural and Biological
Wolt JD (2004) A meta-evaluation of nitrapyrin agronomic and environmental effectiveness with emphasis on corn production in the Midwestern USA. Nutrient
Engineers Annual International Meeting.
Ju XT, Liu XJ, Zhang FS, Roelcke M (2004) Nitrogen fertilization, soil nitrate accumulation, and policy recommendations in several agricultural regions of China.
Ambio: A Journal of the Human Environment, 33, 300–305.
Ju XT, Xing GX, Chen XP et al. (2009) Reducing environmental risk by improving N
management in intensive Chinese agricultural systems. Proceedings of the National
Academy of Sciences of the United States of America, 106, 3041–3046.
Cycling In Agroecosystems, 69, 23–41.
Zerulla W, Barth T, Dressel J et al. (2001) 3, 4-Dimethylpyrazole phosphate (DMPP) –
a new nitrification inhibitor for agriculture and horticulture – an introduction. Biology and Fertility of Soils, 34, 79–84.
Zhang W, Dou Z, He P et al. (2013) New technologies reduce greenhouse gas emissions from nitrogenous fertilizer in China. Proceedings of the National Academy of
Sciences of the United States of America, 110, 8375–8380.
Kim DG, Giltrap D, Saggar S, Palmada T, Berben P, Drysdale D (2012) Factors controlling disappearance of nitrification inhibitor, dicyandiamide (DCD) in a grazed pasture soil
in Manawatu. Proceedings of the Workshop “Advanced Nutrient Management:
Gains from the Past–Goals for the Future”. pp. 7–9.
Kusiima JM, Powers SE (2010) Monetary value of the environmental and health externalities associated with production of ethanol from biomass feedstocks. Energy
Policy, 38, 2785–2796.
Le C, Zha Y, Li Y, Sun D, Lu H, Yin B (2010) Eutrophication of lake waters in China:
cost, causes, and control. Environmental Management, 45, 662–668.
Li SX, Wang ZH, Stewart BA (2013) Responses of crop plants to ammonium and
nitrate N. In: Advances in Agronomy (ed Donald L), pp. 205–397. Academic Press,
New York.
Liu LL, Greaver TL (2009) A review of nitrogen enrichment effects on three biogenic
GHGs: the CO2 sink may be largely offset by stimulated N2O and CH4 emission.
Ecology Letters, 12, 1103–1117.
Liu J, You L, Amini M, Obersteiner M, Herrero M, Zehnder AJ, Yang H (2010) A
high-resolution assessment on global nitrogen flows in cropland. Proceedings of the
National Academy of Sciences of the United States of America, 107, 8035–8040.
MPI (2013) News: DCD Suspension Supported. Ministry for Primary Industries,
Wellington, New Zealand. Available at: http://www.mpi.govt.nz/newsresources/news/dcd-suspension-supported (accessed 24 January 2013).
Rockstr€
om J, Steffen W, Noone K et al. (2009) A safe operating space for humanity.
Nature, 461, 472–475.
Smil V (1999) Nitrogen in crop production: an account of global flows. Global Biogeochemical Cycles, 13, 647–662.
Sterne JAC, Egger M, Smith GD (2001) Systematic reviews in health care – investigating and dealing with publication and other biases in meta-analysis. British Medical
Journal, 323, 101–105.
Sutton MA, Howard CM, Erisman JW et al. (2011) The European Nitrogen Assessment: Sources, Effects and Policy Perspectives. Cambridge University Press,
Cambridge, UK.
USDA (2012) Crop Production. The National Agricultural Statistics Service (NASS),
Agricultural Statistics Board. United States Department of Agriculture, Washington, DC.
USDA (2013) Feed Grains Data Yearbook Tables: Table 12-Corn: Cash Prices at Principal
Markets. Economic Research Service, United States Department of Agriculture,
Washington, DC. Available at: http://www.ers.usda.gov/datafiles/Feed_Grains_
Yearbook_Tables/Domestic_and_International_Prices/FGYearbookTable 12.xls
(accessed 15 January 2013).
Van Grinsven HJ, Holland M, Jacobsen BH, Klimont Z, Sutton MA, Jaap Willems W
(2013) Costs and benefits of nitrogen for Europe and implications for mitigation.
Environmental Science & Technology, 47, 3571–3579.
Vitousek PM, Menge DNL, Reed SC, Cleveland CC (2013) Biological nitrogen fixation: rates, patterns and ecological controls in terrestrial ecosystems. Philosophical Transactions of the Royal Society of London. Series B, Biological Sciences,
368, 20130119.
© 2014 John Wiley & Sons Ltd, Global Change Biology, 21, 1249–1257
Supporting Information
Additional Supporting Information may be found in the
online version of this article:
Figure S1. Global distribution of study sites included in this
synthesis. Red triangles represent agriculture sites and purple triangles represent pasture sites.
Table S1. Publications and observations numbers of each
variable for this meta-analysis.
Table S2. Response ratios of variables for soil chemistry
with 95% bootstrap CI.
Table S3. Response ratios of variables for soil N leaching
with 95% bootstrap CI.
Table S4. Response ratios of variables for GHG emission
with 95% bootstrap CI.
Table S5. Response ratios of variables for air pollutant emission with 95% bootstrap CI.
Table S6. Response ratios of variables for plant productivity
with 95% bootstrap CI.
Table S7. Response ratios of variables for plant nitrogen use
efficiency with 95% bootstrap CI. In Table S2–S7, number of
studies for each category contained is shown in following
parentheses. Numbers highlighted by bolded font indicate
that the 95% bootstrap CI does not overlap 1.
Table S8. Effects of NI on between-group heterogeneity
(Qb) for each assessed variable. Numbers highlighted with
bolded font indicate that the response ratios are different
among the subgroups.
Table S9. Results of publication bias tested by Egger’s
regression test. P-values < 0.05 indicated publication bias
existed.
Data S1. Source references of all datasets in current
syntheses.