Mass loss and N content of the shoot litter material showed that litter-N mass was already depleted due to microbial utilisation by day 66. In the present study the increasing soil NH4+ concentrations over time appear to have come from the soil-N pool since this effect was
from a control treatment, 60 d after surface application of an oats-legume cover crop (C: N, 30) at different moisture levels and indicated that lower NH4+ concentrations from the cover
crop was due to N immobilisation in the initial 5 d.
Brunetto et al. (2011) tracked the decomposition of 15N-labelled ryegrass and clover in vineyards over 16 weeks and found 45% mass loss after 56 d of incubation when using
surface-placed litterbags. Mass loss in the present study on day 66 from the ‘SL’ treatment was similar (46%) to Brunetto et al. (2011) however the values were almost double (82%) for the ‘CL’ treatment on the same day. This might be because in the ‘CL’ treatment, presence of undisturbed plant roots and associated microbes might have aided in decomposition of the added litter. However, in the present study, soil analyses was not performed during the initial period (10–20 d after treatment), when the microbial activity was assumed to be maximum.
Tutua et al. (2002) surface-placed a mixture of ryegrass and clover in a New Zealand apple orchard and found that about 60% of the initial mass was lost after 90 d from the placement of litterbags; the faster rate of decomposition was attributed to irrigation and soil burial of the litterbags. The ‘SR0’ treatment had similar values (50% at day 66) compared to
the results of Tutua et al. (2002) but the value of 53% on day 139 was presumably due to the recalcitrant biochemical composition of the root material (Glasener et al. 2002; Wang et al. 2010). Tutua et al. (2002) reported higher decomposition rates from the buried litterbags than the surface-placed litterbags. Mass loss from the ‘SR2’ and ‘SR0’ treatments in this study were
similar by day 66 but it increased significantly to 84% by day139 – the maximum loss compared to all the treatments thus showing the effect of incorporation, similar to the results of Tutua et al. (2002). However, the observed increase in the litter mass of the shoot
treatments at day 139 in the current might be an artefact due to new herbage growing through the mesh of the litterbags in the ‘CL’ and ‘SL’ treatments.
Recent studies using litterbags (Sanaullah et al. 2010; Wang et al. 2010) have reported changes in the C and N composition of the litter over time. Wang et al. (2010) investigated the decomposition of dried leaf litter of forest tree species in litterbags and found that the C content of the litter decreased over 12 months while the N content increased over the same period while Sanaullah et al. (2010) reported reductions in the C and N contents of leaf litter of ryegrass over 11 months. Kriauciuniene et al. (2008) reported similar results from root and shoots of rape (Brassica napus L.), winter wheat (Triticum aestivum L.) and red clover (Trifolium pratense L.) with maximum reduction during 33–63 weeks. The present study showed reductions in C and N contents from both shoot and root treatments with an exception that the N content of root material increased over time, similar to Wang et al. (2010). This may have been due to N immobilisation as reported by Wang et al. (2010).
Unlike the current study, others have found that surface application of plant material increases the CO2 emissions (Geisseler and Horwath 2011). However, in this current study the
CO2 emissions peaked at days 6 and 7 irrespectively of the soil temperature and soil water
contents, which coincided with peak N2O emissions. This may have been due to higher
microbial activity in the initial phase of treatment application, when soil analyses were not performed. Flessa et al. (2002) reported elevated CO2 emissions immediately after grass
mulch application and concluded that CO2 was emitted from the combination of soil and plant
respiration and activity of decomposer organisms.
8.5 Conclusion
Surface decomposition of ryegrass shoots and roots stimulated N2O emissions with
maximum emissions 5–10 d after treatment application with emissions continuing for a period of 49 d. The 15N analyses showed that litter-N made a significant contribution to the N2O flux
and approximately 70% of the total N2O in the ‘CL’ treatment and 40–50% in the ‘SL’
treatment originated from the litter in the surface-placed shoot litter (‘CL’ and ‘SL’) treatments. The elevated emissions in the shoot litter treatments are attributed to the lower C: N ratio and rich biochemical composition of the ryegrass. The N2O emissions most likely
resulted from a coupling of nitrification and denitrification of the litter-derived-N. An emission factor of 0.9 ± 0.2% was calculated for the ‘shoot only’ (CL) treatment which is similar to the EF values in the lab study (Chapter 6) and to the default EF of 1% stipulated by the IPCC for crop residues. This treatment (CL) also most closely resembles the effect of in
situ litter-fall onto grazed pasture. The implications of a 0.9% EF for litter-fall are discussed
in Chapter 9. Litter treatments did not affect CO2 emissions. Dry matter loss from the
litterbags ranged from 46–82% for the shoot treatments and 50–57% for the root treatments. The C and N contents of the litter from litterbags decreased over time. Investigation using pasture litter of varying biochemical composition is warranted to further consider the impacts of litter rates and the microbial processes responsible for the N2O emissions observed