Effect of dicyandiamide on nitrous oxide flux following return of animal excreta to grassland

Soil Bid.

Peqpmon PII: soo38-cnl7(w)oooseo

SHORT

Biochem.

Vol. 29, No. S/10. pp. 1575-1578,1997 Q 1997ElsevierScience Ltd. All rights reserved Printed in Great Britain 0038-0717/97$17.00+ 0.00

COMMUNICATION

EFFECT OF DICYANDIAMIDE ON NITROUS OXIDE FLUX FOLLOWING RETURN OF ANIMAL EXCRETA TO GRASSLAND J. C. WILLIAMSON’*

and S. C. JARVIS’

‘Manaaki Whenua-Landcare Research, Private Bag 3127, Hamilton, New Zealand and IInstitute of Grassland and Environmental Research, North Wyke, Okehampton EX20 2SB U.K. (Accepted 24 January 1997

concern regarding the contribution of agricultural practices on the gaseous emissions that have been implicated in global climate change

There is growing

and ozone depletion. For example, fertilizer N is regarded as one of the most important sources of

anthropogenic nitrous oxide (N20) emissions (Delgado and Mosier, 1996), and grasslands contribute about 10% of the global atmospheric N20 flux (Bouwman et al., 1993). In addition to N input from grazing animals, grasslands often receive slurry from housed animals and dairy shed waste water. de Klein and van Logtestijn (1994) showed that urine application significantly increased N20 flux and calculated that N20-N loss was 8-16% of the applied urine-N in grassland soils. Use of nitrification inhibitors during autumn and winter can reduce the amount of nitrate leached from pasture soil receiving dairy waste water (Williamson et al., 1996). However, it is uncertain from the literature what effects nitrification inhibitors have on N20 flux, particularly under a manurial-N regime. Skiba et al. (1993) found that DCD inhibited N20 flux following mineral-N fertilizer application by up to 40%. Our objective was to determine the effect of DCD on N20 flux from a grassland receiving animal excretal N. Eight plots, each 6 m2, were sited on grassland over a poorly drained silty clay loam classified as a Halstow series (U.K.) non-calcareous Pelosol (USDA Ochrept), with a clay content of 36% in the top 10 cm and a CEC of 29 cmol kg-‘. The site was located at North Wyke, Devon, U.K. Additional soil characteristics are given in Table 1. The plots were randomly allocated one of eight treatments: dairy cow dung 4 DCD, urine $- DCD, dairy waste water k DCD and KN03 f DCD. The dairy wastes and KN03 were each applied in November 1995 at a rate of 60 kg N ha-’ and diluted with tap water *Author for correspondence.

)

prior to application to ensure similar hydraulic loadings for all treatments (3.3 mm for dung, urine and KN03 and 5.8 mm for waste water). DCD was applied at 7.9 kg N ha-’ and mixed with the diluted wastes or KNOX. A nearby untreated area was designated the control. The soil moisture deficit of the site was restored to field capacity following heavy rain on the fourth day after application of the treatments [Fig. l(a)]. Dung and urine samples were collected fresh from a milking herd grazed on a ryegrass-clover sward. Waste water was obtained from a below ground reception pit which was always covered. The contents of the pit were not stirred or filtered. Waste water was generated from washing the milking parlour and concrete waiting areas as well as rain water runoff from yards and buildings. It did not include any bedding material. Compositions of the three excretal types are presented in Table 1. The application of KNO, was included to provide a defined NO? source to identify N20 flux produced by denitrification rather than nitrification. However, the effect of KN03 addition on NH: exchange behaviour, and hence possible stimulation of nitrification from NH: release, was not known and was investigated by measuring the concentration of water-extractable NH: in soil before and after KNO3 addition, in a laboratory study. Waterextractable NH: was assumed to be bioavailable for nitrification. Duplicate samples of freshly collected soil from the trial site received KN03 at the same rate and same hydraulic loading as in the field trial. Sub-samples (15 g) were taken from replicate treatments and a control at 0.3 h and 1.0 h following amendment: soils were extracted in deionized water (1:2 soikextractant) and filtrates analysed for automated calorimetry ammonium-N by (Blakemore et al., 1987). Two static chambers (40 cm dia x 30 cm height) per plot were used for the in-situ measurement of

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Table I. Composition of soil and excretal types Type

Dry matter (%)

Total N (%‘)

ND 12 ND ND

0.26 2.88 0.84 0.02

SoiP Dung Urine Waste water

Total C (“4~‘)

NH: (pg N cm-‘)

NOT (pg N cm-“)

pH(Hz0)

ND ND 180 102

ND ND 0 0

5.1 ND 8.0 8.3

2.6 43.0 1.4 ND

“w/w oven-dry basis for soil and dung: w/v for urine and waste water hO-IO cm depth. ND, not determined.

NzO flux. After sealing the chambers for 1 h, duplicate 60 ml samples of headspace gas were withdrawn from three-way valves fitted to the lid of each chamber and the accumulated N20 determined by gas chromatography (Allen et al., 1996). Flux

0

10 20 30 Time (days) relative to Start Oftdal

T

2000,

40

n)

measurements were undertaken nine times over 37 d. Values for the four N20 concentrations (pg NzO-N rne3) per treatment were averaged and ambient N20 subtracted, corrected for temperature and a chamber volume:surface area ratio of 1:4 to provide a mean flux per treatment (fig NzON m-’ h-l). Soil temperature (10 cm) and rainfall data during the trial are presented in Fig. l(a). All results are on an oven-dry soil (105°C) weight basis. Significant differences are expressed at P < 0.05, unless otherwise stated. Analysis of variance and multiple pairwise comparisons were performed using the MINITAB statistical package. Mean N20 flux from control soil was 40 pg NzON m-* h-’ and did not vary significantly with time. Nitrogen addition, regardless of substrate, led to increased N20 fluxes; temporal patterns of N20 flux were similar for the different substrates [Fig. l(b-e)]. N20 fluxes from the substrates varied considerably in magnitude, with KN03 stimulating the greatest flux [Fig. l(b); maximum of 164Opg N20-N m-* h-l], followed in descending order by urine [Fig. l(c)], waste water [Fig. l(d)] and dung [Fig. l(e); maximum of 160 pg N20-N m-* h-l]. Temporal patterns followed changes in precipitation (and therefore soil water content) and soil and headspace temperatures [Fig. l(a)]: this was consistent with the conclusions of Allen et al. (1996). Fluxes from the urine-treated soil peaked within 10 d of application and had mostly subsided after 30 d [Fig. l(c)]: this was also consistent with the results of Allen et al. (1996) for the same soil type. However, the effects of dung on soil [Fig. l(e)] were significantly greater in our study because dung was applied with water in contrast to undiluted pats used by Allen et a/. (1996). The difference in N20 flux between urine and dung treatments probably reflected the bioavailability of N. Urine contains 65-90% urea (Lockyer and Whitehead, 1990) which rapidly hydrolyses to ammonium, whereas the organic forms of N in dung are more recalcitrant and slower to fuel nitrification and denitrification processes. DCD significantly reduced NzO flux from the KN03 treatment between 6 and 30 d after application [Fig. l(b)]: summing the N20 emitted over the six one-hourly measurements taken over this period, 5000 pg N20-N me2 were emitted from the KNOs treatment and 1200 pg N20-N m-* from the

:,I #iliz&_ 1,

a

T

30*j JLikh-d 1 0

T

TT

10

e)

-r

20

30

40

Time (days) relative to start of trial

Fig. 1. Time course of: (a) soil temperature (---), peak rainfall events (- - - -) and chamber headspace temperature (0) during the field trial; (b) N20 flux from applied KNOX (+) and KNO, + DCD (4); (c) N20 flux from applied dairy cow urine (+) and urine + DCD (+); (d) N20 flux from applied dairy waste water (+) and waste water + DCD (+); and (e) N20 flux from applied dairy cow dung (+) and dung + DCD (+). Dairy wastes and KNOs were applied at 60 kg N ha-’ to grassland: DCD was applied at 7.9 kgN ha-‘. All flux values are means of four replicates; vertical bars are SE of means. Flux from a control area (0) is presented in (b)-(e).

Short Communication

KNO, + DCD-treatment. We had assumed that DCD would have little or no effect on NzO flux from KN03 application since the overriding contribution to NzO production would be from denitrification. Skiba et al. (1993) reported a decrease in NzO emission by DCD (applied at 10 kg ha-‘) after applying KN03 at 100 kg N ha-‘. We found that adding KN03 at an equivalent rate of 60 kg N ha-’ to untreated soil from the field trial increased water-extractable NH: within 0.3 h by 6.5 pg N g-’ soil, a ninefold increase compared to soil that did not receive KN03, which was sufficient to account for observed difference in field fluxes. Similarly, Skiba er al. (1993) found a threefold increase in soil available NH; immediately after an application of KNOX. We suggest that the addition of KN03 resulted in a rapid displacement of NH; from the cation exchange sites of this silty clay loam, which became directly available for nitrification. This suggests that NzO flux from the KN03 + DCDtreatment was derived solely from denitrification, whereas the NzO flux from the KNOrtreatment was derived from denitrification of added KN03 and nitrification and subsequent denitrification of the displaced NH:. This demonstrates the difficulty in assessing the relative contributions of nitrification and denitrification to NzO flux using DCD with KN03. DCD was effective in significantly reducing NzO flux from the urine treatment between 6 and 21 d [Fig. l(c)]: summing the NzO emitted over the five l-h measurements taken over this period, 2220 pg NzO-N me2 were emitted from the urine treatment and 580 pg N20-Nm-* from the urine + DCD-treatment. Dung + DCD-treated plots emitted less N20 than dung-treated plots on average [Fig. l(e)], but the differences were non-significant, except on day 16 (P < 0.10). There was no significant effect of DCD when added to waste water [Fig. l(d)] which was surprising, given that the main N components were derived from urine and dung. The reason for the nil-response is not known, but there may have been an effect of storage. There was a substantial flux from waste water at days 6 and 7 (ca. 300 pg N20-N me2 h-l), and since there was no NOj detected in waste water (Table 1), then nitrification must have occurred for N20 to have been generated. However, nitrification can occur through both autotrophic and heterotrophic pathways, but DCD only inhibits autotrophic nitrifying bacteria (Kutuzova and Tribis, 1993). This raises the possibility that the chemistry of stored waste water favoured heterotrophic nitrification pathways over chemoautotrophic ones, when applied to land. Storage of waste water often results in the accumulation of volatile fatty acids which, when applied to soil, cause N-immobilization by heterotrophic microorganisms (Kirchman and Lundvall, 1993). Under such conditions, auto-

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trophic nitrifying bacteria may prove to be poor competitors for N. A shift to heterotrophic nitrification would explain the nil-response to DCD seen in the waste water treatment. The likely mechanisms by which DCD affected N20 flux were: (i) reduced N20 production by direct inhibition of nitrification, and (ii) reduced N20 production from denitrification through substrate (NOT) limitation. High soil NO; concentrations can inhibit the reduction of N20 to Nz (Weier er al., 1993), and therefore DCD may reduce N20 emission indirectly by reducing soil NO; concentration. Although this trial did not extend beyond 37 d. studies by Allen et al. (1996), conducted at the same site in November 1993 under similar soil moisture conditions (approximately field capacity) showed, during an 80-d trial, that N20 fluxes from urine and dung had mostly ceased by 30 d. The half-life of DCD under prevailing soil temperatures [mean 10°C: Fig. l(a)] would ensure that recovery of nitrification would be delayed until the onset of increased plant N-uptake (Williamson et al., 1996), precluding subsequent, substantial emissions. Dicyandiamide was effective in reducing N20 flux rates from grassland receiving fresh dairy cow urine and, to a lesser extent, dung, but not stored waste water. Future work needs to address how storage of waste water affects the inhibitory capacity of DCD, particularly as DCD was found to inhibit nitrification of a fresh waste water application effectively (Williamson et al., 1996). Acknon,ledgements-Julie Williamson was supported by the Ministry of Research, Science and Technology and Foundation for Research, Science and Technology, New Zealand. The work was partly funded by the Ministry of Agriculture, Fisheries and Food, London. We are grateful to P. Owen. E. Dixon and E. Williams for technical assistance, to R. Lovell. A. Bristow and S. Yamulki for technical advice and to K. Tyson for meteorological data.

REFERENCES Allen A. G., Jarvis S. C. and Headon D. M. (1996) Nitrous oxide emissions from soils due to inputs of nitrogen from excreta return by livestock on grazed grassland in the U.K.. Soil Biology & Biochemistry 28, 597-607.

Blakemore, L. C., Searle, P. L. and Daly, B. K. (1987) Methods for chemical analysis of soils. New Zealand Soil Bureau Scientific Report, No. 80. 103 pp. Bouwman A. F., Fung I., Matthews E. and John J. (1993) Global analysis of the potential for NzO production in natural soils. Global Biogeochemical Cycles I, 557-597. Delgado J. A. and Mosier A. R. (1996) Mitigation alternatives to decrease nitrous oxide emissions and urea-nitrogen loss and their effect on methane flux. Journal of Environmental Quality 25, 1105-l I 11. Kirchman H. and Lundvall A. (1993) Relationship between N immobilization and volatile fatty acids in soil after application of pig and cattle slurry. Biology and Fertility of Soils 15, 161-164.

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de Klein C. A. M. and van Logtestijn R. S. P. (1994) Denitritication and NzO emission from urine-affected grassland soil. Plant and Soil 163, 235-242. Kutuzova R. S. and Tribis Zh. M. (1993) Nitrification inhibitors and stage II of the autotrophic process. Eurasian Soil Science 25, 68-84.

Lockyer D. R. and Whitehead D. C. (1990) Volatilization of ammonia from cattle urine applied to grassland. Soil Biology & Biochemistry 22, 1137-l 142.

Skiba U., Smith K. A. and Fowler D. (1993) Nitrification and denitrification as sources of nitric oxide and nitrous

oxide in a sandy loam soil. Soil Biology & Biochemistry 25, 1527-1536.

Weier K. L., L. (1993) oxide ratio and nitrate.

Doran J. W., Power J. F. and Walters D. Denitrification and the dinitrogen/nitrous as affected by soil water, available carbon, Soil Science Society of America Journal 57,

66-72.

Williamson J. C., Menneer J. C. and Torrens R. S. (1996) Impact of dicyandiamide on the internal nitrogen cycle of a volcanic, silt loam soil receiving effluent. Applied Soil Ecology 4, 39-48.