«International 17 Workshop th Nitrogen The was jointly organised by Teagasc and AFBI Printed by Print Depot Suggested citation Authors, 2012. Title ...»
N fertilization and diazotrophic bacteria inoculation in sugarcane for bioenergy production 339 H. Cantarella, Z. F. Montezano, G. J. C. Gava, R. Rossetto, A. C. Vitti, V. P. Vargas, J. Soares, C. A. Oliveira, H. A. W. Joris, O. T. Kölln, F. L. T. Dias, S. Urquiaga N2O emissions from Radiata Pine, Douglas fir and Beech forest stands in the Basque Country 341 I. Barrena, J. M. Estavillo, M. Duñabeitia, P. Merino, C. González-Murua, S. Menendez
Nitrogen use efficiency improvement in heavy-pig production in Northern Italy 350 G. Della Casa, M. T. Pacchioli, R. Marchetti
Prediction of nitrous oxide emissions from Irish arable lands using the ECOSSE model 358 M. I. Khalil, J. Smith, M. Abdalla, P. O'Brien, P. Smith, C. Mueller, M. Richards
Temporal dynamics of soil N mineralization during an oilseed rape (Brassica napus l.) growth cycle in one season´s growth under humid mediterranean conditions 368 N. Villar, P. Gallejones, A. Castellón, G. Besga, A. Aizpurua
The Environmental virtual Observatory (EvO): can cloud-based modelling provide new understanding of nutrient cycling processes from catchment to national scale? 372 S. Greene, P. J. Jones, J. Freer, N. O'Doni, J. Bloomfield, S. Reaney, C. J. A. McLeod
use of chemical amendment of dairy cattle slurry to reduce phosphorus losses from dairy cattle slurry while allowing land spreading of slurry to meet nitrogen requirements. 379 R. Brennan, M. Healy, G. Lanigan, O. Fenton
using the Eurotate_N crop model to optimize nitrogen fertilization in potato crop 381 L. Olasolo, N. Vázquez Garcia, M. L. Suso, A. Pardo S.3.IP Global Perspectives on Nitrogen and Food Security Invited Presentations
The product carbon footprint of milk from pasture- and confinement-based dairy farming 400 S. Philipp, T. Biegemann, M. Kamper, R. Loges, F. Taube S.3.SP Global Perspectives on Nitrogen and Food Security Short Presentations
Integrated assessment of nutrient management options in the food chain of China 406 L. Ma, F. H. Wang, W. F. Zhang, W. Q. Ma, G. Velthof, W. Qin, O. Oenema, F. Zhang S.3.P Global Perspectives on Nitrogen and Food Security - Posters
Effect of nitrogen fertilizer application timing on yield of winter wheat in Ireland 411 A. Efretuei, M. Gooding, E. White, R. Hackett, J. Spink
how does sheep grazing affect the greenhouse gas balance of a grazed steppe ecosystem? 415 S. Philipp, B. Wolf, M. Wiesmeier, U. Dickhofer, H. Wan, M. Gierus, K. Butterbach-Bahl, A. Susenbeth, F. Taube Influence of different nitrogen fertilizers on forage maize yield an quality 417 M. I. García Pomar, D. Báez Bernal, A. Louro López, J. Castro
Soilless cultivation of vegetables in The Netherlands to reduce nitrogen emissions 423 D. H. Janjo, E. van Os, M. Blind, V. John
Extension & Knowledge Transfer; Effective Partnerships for Timely Impact 432 Q. Ketterings, K. Czymmek S.4.OP Knowledge Transfer - Long Presentations
Estimating the effect of mitigation methods on multiple environmental pollutants 439 P. Newell Price, D. Harris, D. R. Chadwick, S. G. Anthony, R. D. Gooday, M. Taylor, J. R. Williams, B. J. Chambers S.4.SP Knowledge Transfer - Short Presentations Achieving good water quality status in intensive animal production areas: a lIFE+ project 441 E. Bortolazzo, M. Ligabue, M. T. Pacchioli, P. Mantovi Nitirsoil: a new N-model to estimate monthly nitrogen soil balance in irrigated agriculture. 443 J. M. de Paz, C. Ramos, F. Visconti
"Reliquat virtuel" : a new decision support tool to predict the soil inorganic N pool 448 N. Damay, C. Le Roux, J. Gaillard, J. MacHet
Quantitative evaluation of hot water extractable organic matter of organic farm soils in Japan by measurement of chemical oxygen demand with inexpensive chemicals and equipment 474 K. Kanazawa, S. Takahashi, M. Komada, N. Kato Satellite data potential for assessing potato crop nitrogen status at a specific field scale 476 G. Jean-Pierre, L. van Den Wyngaert, D. Buffet, A. Leonard, P. Defourny
using canopy reflectance to determine appropriate rate of topdress N in potatoes 484 F. K. van Evert, D. D. A. van Der Schans, J. T. Malda, W. van Den Berg, W. van Geel, J. N. Jukema
Advances in understanding nitrogen flows and transformations where is the missing nitrogen?
Müller, C.a,b and Clough, T.J.c a Department of Plant Ecology, Justus-Liebig-University Giessen, Germany b School of Biology and Environmental Science, University College Dublin, Ireland c Faculty of Agriculture and Life Sciences, Lincoln University, New Zealand
Nitrogen transformations and balances: gaps and research pathways
1. Introduction Anthropogenically generated reactive nitrogen (N) cascades throughout the global environment (Galloway and Cowling, 2002). This reactive N may be lost from ecosystems via leaching, as nitrate (NO3-), or in gaseous forms such as ammonia and nitrous oxide (N2O). These N losses are of significant importance both economically and environmentally. Despite more than 150 years of N cycling research by well known scientists such as Liebig (van der Ploeg et al., 1999), Bausingault (Aulie, 1970), Winogradsky (Ackert, 2006) and others there are still significant questions to be addressed with respect to N transformations and losses from terrestrial ecosystems.
Plant effects Plant-soil interactions are an area of research attracting increasing interest with respect to N cycling.
They are predominantly governed by interactions between the carbon (C) and N cycles. The microbial loop in soil is driven by recent plant C inputs (e.g. rhizodeposition) which influences the microbial activity in soils and ecosystem N availability is controlled by the initial chemical composition and litter N concentrations (Manzoni et al., 2008; Parton et al., 2007). In the rhizosphere, potential N transformations such as denitrification decrease rapidly in the first few millimetres of roots (Beauchamp et al., 1989). Thus soil aggregate-microsite reactions and the influence of roots are likely to cause small-scale variations of substrate availability and environmental regulators and prompt for a diverse reaction of the microbial community (Barnard et al., 2005).
Nitrogen cycling is closely associated with plant productivity and factors affecting it. For example, conditions that favour plant C assimilation may also enhance rhizodeposition and subsequently alter microbial community composition and function with subsequent effects on N cycling. Conversely, a higher demand for N by plants, that may occur as a result of environmental change, e.g. favourable growth temperatures, increases competition for N between plants and microbes, potentially affecting microbial community structures and subsequent N transformations in the soil (Zak et al., 2003). However, only few studies have examined the links between the N cycle, plant activity and associated changes in microbial form and function that affect N transformations and fluxes. In Nlimited systems, the gross N transformation rates and not the sizes of the soil N pools govern the availability of N for plants and microbes (Rastetter et al., 1997).
Plants utilise mineral N (e.g. NH4+ or NO3-) and low molecular organic N compounds. The latter enter the soil via rhizodeposition (amino acids etc.) or are made available following microbial mineralization of organic material. Organic N (dissolved organic nitrogen, DON) uptake is reported to be more significant under conditions of N limitation and low pH (Jones et al., 2004). Despite the significance of DON as a loss pathway in agricultural ecosystems (Van Kessel et al., 2009) current knowledge of the organic N forms utilized by plants, with respect to DON dynamics over time and
Nitrogen Workshop 2012
space in the soil, and their rates of production and utilization is still limited. Similarly, the role DON plays in gaseous N loss pathways is also under researched. A recent review indicates that nitrosation reactions may be an important loss pathway for gaseous N (Spott et al., 2011).
Rhizodeposition stimulates microbial biomass growth and N turnover (Knops et al., 2002), which further stimulates the growth of predating organisms such as protozoa and nematodes (Osler and Sommerkorn, 2007; Sanderman and Amundson, 2003). Thus soil organic matter (SOM) is mineralized and decomposition of SOM may even be accelerated (priming). So far a quantitative evaluation of the various and simultaneously occurring N transformations, that lead to mineralization and immobilization of N in the rhizosphere is lacking. Can plants actively control exudation to create conditions in the rhizosphere to maximize the availability of N? Do plants actively produce compounds that deter non-beneficial organisms/pathogens directly or indirectly via relationships along trophic levels that in turn affect N cycling? Rhizodeposition of C compounds can directly influence N cycling, for instance enhancing N2O emissions in the presence of an N substrate (Uchida et al., 2011). But relatively little is known about the forms and rates of C released.
N loss from ecosystems – the importance of denitrification A major loss pathway for N to be released into the environment is via leaching as NO3-. To be rendered environmentally benign this NO3- must be reduced to a non-reactive form, dinitrogen (N2).
Three major biological pathways of NO3- reduction are known: i) assimilatory NO3- reduction into biomass, ii) dissimilatory NO3- reduction to NH4+ (DNRA) and iii) dissimilatory NO3- reduction to N2 (denitrification) (Burger and Jackson, 2004; Robertson and Kuenen, 1990). DNRA may outcompete denitrification under conditions when electron acceptors are limited and provides an energetically favourable alternative to denitrification (Rütting et al., 2011; Tiedje, 1988). However, further research is still required to understand the importance of DNRA in terrestrial systems and to perfect methods for studying the process (Rütting et al., 2011).
Denitrification is a key transformation process in soils with adverse and beneficial roles, since it impairs N use efficiency of agricultural crops, is both a source and sink for N2O, and lowers the potential for NO3- leaching to aquatic systems (Davidson and Seitzinger, 2006). Denitrification is a heterotrophic process performed by facultative anaerobic organisms that utilize various C substrates as electron donors (Beauchamp et al., 1989). In aquatic systems, autotrophic denitrification may also occur in the presence of inorganic electron donors like sulphides or ferrous iron (Clément et al., 2005; Knowles, 1982).
The reductases involved in denitrification are well recognized. The membrane bound nitrate reductase (Nar) is the first in the denitrification sequence and occurs in very diverse microbial communities including a, b, g, and e proteobacteria, gram positive bacteria and archaea (Philippot,
2005) while the periplasmic NO3- reductase (Nap) occurs only in gram-negative bacteria (Philippot, 2005). The key enzyme and key precursor for gaseous N emissions in the denitrification sequence is nitrite reductase which is encoded by either nirS, a cytochrome cd1 containing gene, or nirK, a copper containing gene. The two nitrite reductases provide a functional marker for the diversity of denitrifying bacteria (Braker et al., 2000). The genes norB and nosZ encode the NO and N2O reductase respectively (Groffman et al., 2006) and it is the dynamics of these two reductases in relationship to substrate and environmental regulators and in particular the balance between nar/nir and nos which controls the net production and release of N2O. Philippot et al. (Philippot et al.,
Nitrogen Workshop 2012
2011) showed that denitrifiers with and without nosZ encoding can co-exist and that higher N2O/N2 ratios can be related to the proportions of bacteria with and without nosZ encoding. Differential encoding may also be related to soil pH which may differ in distinct niches (Philippot et al., 2009).
This highlights the fact that the observed overall denitrification dynamic is governed by the sum of the individual dynamics of co-existing microbes, possibly living in different niches (Philippot et al.,
2011) and/or soil aggregates (Miller et al., 2009). Soil aggregate size also influences the rate of denitrification despite denitrification activity and denitrifier abundance not being associated at the aggregate level (Miller et al., 2009). Moreover, aggregation affects diffusive exchange of N2O and O2 between denitrifying microsite and inter-aggregate pores, which affects both, total denitrification (Sexstone et al., 1985) and the product ratio of denitrificication (Arah and Smith, 1989).