Greenhouse-gas abatement on Australian dairy farms: what are the options?
L. M. Garnett
A * and R. J. Eckard B
A
B
Abstract
The Australian dairy industry contributes significantly to the rural economy, but must reduce its greenhouse-gas emissions to remain competitive in a global market that is starting to prioritise a low carbon footprint. Demand for improved environmental, social and governance performance from supply chains creates an imperative for research to deliver options for farmers to make reductions in their environmental footprint. Given the rapidly evolving nature of greenhouse-gas abatement research, this critical review provides an update on the state of the research relevant to Australian dairy systems and identifies research gaps that must be addressed if there is to be widespread on-farm adoption. Current research suggests that Australian dairy farms could theoretically abate enteric methane by 40–50%, with about another 5–10% reduction in whole-farm greenhouse-gas emissions being possible by flocculating or covering stored effluent. Fertiliser- and urine-patch management strategies could substantially reduce direct and indirect nitrous oxide emissions, but by variable amounts subject to local conditions. However, few abatement options are currently cost-effective for farmers. Significantly more research investment is required to facilitate the on-farm adoption of strategies, particularly to reduce enteric methane and improve the efficiency of nitrogen cycling. Improved understanding is required of the influences on each strategy’s abatement potential and interactions with economically important traits in grazing systems, the effect of combining abatement strategies, and systems by which strategies can be implemented cost-effectively on farms. The challenge for research is to consider how the implementation of cost-effective abatement options can be refined for grazing dairy systems to maintain the position of Australian dairy in the global market.
Keywords: climate change, manure management, methane, milk production, mitigation, nitrogen, nitrous oxide, sequestration, sustainability.
Introduction
In 2020/2021, the Australian dairy industry was the third-largest rural industry in Australia, generating AU$6.1 billion per year in farmgate value and employing an estimated 33,500 people, either on farm or through dairy processors (Dairy Australia 2023a). Milk products are a nutrient-rich food source for which global demand is growing (Willett et al. 2019; OECD/FAO 2021). Yet, food-production methods must change to reduce their impact on climate change and transgression of planetary boundaries (Willett et al. 2019; Clark et al. 2020).
The dairy industry globally is facing increasing pressure to reduce greenhouse-gas (GHG) emissions from their farming systems, driven not only by national and international targets (e.g. the COP21 Paris Agreement) but also by their supply chains setting targets to align with the Paris Agreement. Given that pre-farm gate emissions make up more than 90% of the total lifecycle emissions of dairy products (Gerber et al. 2010), a range of multinational corporate milk processors, including Nestle, Unilever and Mars, have set targets to reduce Scope 3 (i.e. complete supply chain) emissions by 42–50% by 2030 (SBTi 2024). Meanwhile, dairy manufacturers and retailers that operate in Australia have variable emission-reduction commitments. Scope 3 targets for 2030 have been set by Fonterra and Woolworths Group, whereas Coles Group has committed to setting a Scope 3 target (SBTi 2024). Other Australian processors have not yet committed to setting Scope 3 targets but those that supply local and international companies with targets will need to match these to maintain market access by 2030. This obligation would either require dairy farmers to reduce their farm GHG emissions to maintain market access or result in companies preferencing products with lower carbon footprints. There is also a clear consumer trend towards plant-based milk substitutes, with supply-chain companies starting to include targets for non-animal protein, driven by environmental and animal-welfare concerns (e.g. Janssen et al. 2016; Sebastiani et al. 2019; Aydar et al. 2020; Xu et al. 2021). The Australian dairy industry is therefore at a critical juncture as sustainability becomes as important as price, quality and safety in supply-chain access. It must understand how to mitigate GHG emissions to maintain access to supply chains in the event that targets are enforced.
The Australian dairy industry contributes ~9 Mt of carbon dioxide equivalents (CO2e) per year, which is ~12% of the country’s agricultural emissions (Australian Government 2023). On a typical grazing-based dairy farm in Australia, mean emissions (±s.d.) are 1.07 (±0.02) kg CO2e per kilogram of fat- and protein-corrected milk solids, with over 100% variation between the highest and lowest emitting farms (Christie et al. 2012, 2018a). Globally, the milk production footprint ranges from ~0.7 to 6 kg CO2e/kg fat- and protein-corrected milk solids, with New Zealand, Uruguay and UK dairy farms having lower average footprints than does Australia’s (Mazzetto et al. 2022). Thus, Australian dairy farms have potential for further emissions reduction, particularly from high emission-intensity farms.
Reaching emission targets will require reductions in all GHG pathways and dairy farm emissions are mostly made up of pollutants with a global-warming potential (GWP) far greater than that of carbon dioxide (CO2), namely, methane (CH4; GWP = 28 CO2e) and nitrous oxide (N2O; GWP = 265 CO2e) (Australian Government 2024). Enteric CH4 makes up 56.8% of total emissions on a typical grazing-based dairy farm in Australia, with the remainder consisting of waste CH4 (9.9%), direct and indirect N2O from animal waste (10.7%) and nitrogen (N) fertiliser (2.8%), and CO2e from fuel and electricity (8.9%), purchased feeds (7.9%) and purchased fertiliser (3.1%) (Christie et al. 2018a).
Whereas supply chains may need to purchase low-emission products, based on a standardised emission-intensity metric (e.g. kg CO2e/kg milk solids), governments and multi-national companies (e.g. Unilever 2021) have also set targets for absolute emission (g CO2e/day) reductions in their Scope 3 emissions. A key vulnerability to social licence and international supply-chain access will be an increasing focus on environmental, social and governance (ESG) performance, which requires reducing absolute emissions (e.g. McRobert et al. 2022) and noting restrictions on farm N inputs across most of Australia’s dairy competitors (Powell et al. 2010), plus targets for absolute CH4 reductions (e.g. Global Methane Pledge, California SB-1383 short-lived climate-pollutants bill, New Zealand Climate Change Response Act). Meanwhile, the Australian dairy industry can demonstrate great gains in emissions intensity through adoption of new technology and more efficient management practices, and milk products compare favourably with other foods on a nutrient-density basis (Doran-Browne et al. 2015; Moate et al. 2016; Lean and Moate 2021; Pethick et al. 2023). Further reductions in emissions intensity allow for industry growth to meet food demand with smaller quantities of GHG emitted with each product unit (Moate et al. 2016; Eckard and Clark 2020). However, emission-intensity reductions do not always result in absolute emissions reduction across an industry or region (Ungerfeld et al. 2022). Where production increases outweigh emission-intensity savings, absolute emissions increase (Eckard et al. 2010; Garnett 2014; Leahy et al. 2020; Beauchemin et al. 2022). Therefore, reduction in both emissions intensity and absolute emissions are required to achieve net-zero emissions without compromising food security (Reisinger et al. 2021).
This review paper focuses on options for absolute emissions and emission-intensity abatement of N and CH4 that support the Australian dairy industry to remain competitive in a global market that prioritises low-emission products. GHG-emission abatement from dairy farms is an evolving area of research, with much progress since previous reviews relevant to Australian dairy systems, for example, Moate et al. (2016), Eckard and Clark (2020), Black et al. (2021) and Lean and Moate (2021). Thus, the aims of this review are to identify GHG abatement options relevant to Australian dairy systems, summarise the state of research on their GHG abatement potential, and identify challenges and knowledge gaps that must be overcome to facilitate their implementation.
Enteric methane
Enteric CH4 usually makes up the largest proportion of GHG emissions emitted from dairy farms globally, including in Australia (Mazzetto et al. 2022), and, so, is crucial to dairy GHG abatement globally (Christie 2019). At least 10 approaches have been taken to reducing enteric CH4 production.
Methane inhibitors and rumen modifiers
Methane is produced in the rumen by a group of archaean micro-organisms called methanogens (McAllister and Newbold 2008). Because Archaea are evolutionarily distinct from other rumen organisms and share a common methanogenesis pathway, they can be specifically targeted without affecting beneficial rumen organisms (Hedderich and Whitman 2006; Moate et al. 2016). Theoretically, therefore, modifying the rumen environment to reduce methanogen populations should abate enteric CH4. Archaea use CO2 and hydrogen to produce CH4; so, modifying rumen processes that result in hydrogen could also reduce CH4 production and could theoretically increase animal productivity, because methanogenesis accounts for ~6–10% of gross energy consumed by the cow (Eckard et al. 2010). However, such benefits are constrained by a need for an alternative non-toxic pathway for hydrogen removal from the rumen (McAllister and Newbold 2008; Eckard et al. 2010) because excessive hydrogen in the rumen can affect fermentation and feed digestion (Janssen 2010; Ungerfeld 2020), potentially decreasing dry-matter intake and, ultimately, milk production (Kjeldsen et al. 2024; Maigaard et al. 2024).
Because most dairy cattle receive a grain supplement in the milking parlour twice per day, Australian dairy farms have more options to feed CH4-mitigating supplements or inhibitors, than do other, more extensive, grazing industries (Eckard et al. 2010; Moate et al. 2016; Eckard and Clark 2020). The Australian dairy industry also has greater potential to reduce enteric CH4 significantly by using current technology than does the New Zealand industry, because more Australian farms have supplemental feeding infrastructure (Eckard and Clark 2020), but has fewer opportunities than in countries where total or partial mixed-ration diets are more prevalent and supplements can be provided in every mouthful of feed. As it is, supplements incur a daily cost to farmers, with no opportunity currently available for offsetting via pricing or subsidies. Thus, research is still needed to deliver lower-cost and effective long-term mitigation of CH4 that is practical in grazing-based systems (Eckard and Clark 2020).
Synthetic CH4 inhibitors are promising for CH4 reduction because they can target specific biochemical pathways and be produced in large quantities (Henderson et al. 2018). The most promising synthetic inhibitor, 3-nitrooxypropanol (3-NOP), blocks the active site of the final enzyme in rumen methanogenesis, nickel enzyme methyl-coenzyme M (Duin et al. 2016). Meta-analysis of the extensive research completed in confined dairy systems showed a 32.7%, 30.9% and 32.6% abatement in CH4 production (from 361 to 226 g/day), yield (16.0–10.5 g CH4/kg dry-matter intake, DMI) and intensity (10.6–6.6 g CH4/kg milk solids) respectively, at an average dose of 70.5 mg 3-NOP/kg DMI (Kebreab et al. 2023). Persistent abatement of CH4 production occurs for up to 15 weeks (van Gastelen et al. 2020; Melgar et al. 2021).
Meta-analyses have found that 3-NOP either increases or has no significant effect on milk fat percentage and there is no significant difference in protein percentage or energy or fat-corrected milk yield (Jayanegara et al. 2018; Kim et al. 2020). Risk of toxicity of 3-NOP to people and ruminants is low (Thiel et al. 2019a, 2019b). Although some studies have reported a trend towards decreased DMI associated with increases in 3-NOP dose (Kim et al. 2020; Melgar et al. 2020), more research is required to determine the underlying mechanism (Kjeldsen et al. 2024). 3-NOP is expected to cost ~AU$0.5 per cow per day (Borrello 2023); so, in the absence of other economic incentives, the lack of production benefit will challenge adoption of 3-NOP (Alvarez-Hess et al. 2019).
Limited research has been completed on the effect of 3-NOP on CH4 production and milk production under grazing-based systems where supplementation of 3-NOP can practically occur only twice-daily, during morning and evening milkings, but not during grazing in between (Costigan et al. 2024). Mixed in a total mixed ration, 3-NOP appears to prevent CH4 production peaks that occur following feeding (Vyas et al. 2016; Melgar et al. 2020). Melgar et al. (2020) found that the greatest CH4 production decrease was within 4 h of feeding, approximately matching the peak CH4 production timing of the control diet, indicating that 3-NOP may have to be in every mouthful of feed for substantial CH4 abatement to occur. When supplementing at a rate of 1.36 g/cow.day of 3-NOP over 12 weeks in a grazing-based dairy system, Costigan et al. (2024) found that, although hourly CH4 production was abated by 28.5% during the 3 h after consumption, average daily abatement was only ~5% less than the control amount of 362 g/day and they found around a 36% increase in daily H2 production. They found no impact of feeding 3-NOP on DMI or milk production. Therefore, if 3-NOP is to be a viable option to abate CH4 in grazing systems, extensive further research is needed to address how 3-NOP can be delivered to cows throughout the day while cows are grazing, such as through a cost-effective slow-release supplement or bolus (Costigan et al. 2024).
Red seaweed species, Asparagopsis taxiformis and A. armata, contain high, but varying, concentrations of bromoform and dibromochloromethane, both of which have been shown to inhibit enteric CH4 production (Machado et al. 2016; Vijn et al. 2020; Eason and Fennessy 2023). Meta-analysis has found an average of 39–42% reduction in CH4 yield (g/kg DMI) from red seaweed species supplementation, with no significant impact on DMI, milk yield or components. However, with few studies specifying the Asparagopsis spp. used in the analysis, or quantifying the rate of active ingredients in the experiment, there was a marked heterogeneity in results (Lean et al. 2021; Sofyan et al. 2022). A reduction in enteric CH4 of up to 98% and 99% occurred in two in vivo experiments where beef cattle were fed A. taxiformis (Kinley et al. 2020; Cowley et al. 2024), although with large increases in H2 production, and persistent abatement has been shown for 21 weeks (Roque et al. 2021). Differences in CH4 abatement are potentially caused by the seaweed dose, delivery medium, bromoform concentrations, dietary fibre content or other dietary formulation interactions (Vijn et al. 2020; Roque et al. 2021; Sofyan et al. 2022; Alvarez-Hess et al. 2023).
Research is yet to demonstrate persistent enteric CH4 abatement from feeding Asparagopsis spp. in grass-based dairy systems without affecting milk production. By feeding an oil formulation of A. armata twice daily in a concentrate mix, Alvarez-Hess et al. (2023) showed that CH4 abatement could occur in a system applicable to feeding commercially on Australian dairy farms. However, the A. armata steeped oil mixed with concentrate also reduced concentrate intake, energy-corrected milk yield and fat and protein yield (Alvarez-Hess et al. 2023). Alvarez-Hess et al. (2024) also found that intake of A. armata–treated concentrate decreased with an increasing bromoform concentration. Other studies have also found decreased DMI and milk yield associated with Asparagopsis spp. inclusion (Roque et al. 2019a; Stefenoni et al. 2021; Alvarez-Hess et al. 2024). Research is required to determine the underlying mechanism of the DMI decline. The cost of AU$0.4 or more per cow per day is another a major limitation to adoption (Peel 2024).
Concerns have also been raised regarding potential carcinogenic and ozone-depleting properties of Asparagopsis spp., as well as bromoform and iodine residues found in milk of cows fed Asparagopsis spp. (US EPA 2002; Eckard and Clark 2020; Lean et al. 2021; Reisinger et al. 2021; Stefenoni et al. 2021). However, although bromoform and dibromochloromethane are animal, and probably human, carcinogens (US EPA 2002), the low rates fed so as to reduce CH4 are unlikely to affect animal health and bromoform residues have not been found at dangerous levels in milk or animal tissue when fed to ruminants at minimum effective CH4 mitigation rates (Glasson et al. 2022; Eason and Fennessy 2023). Although iodine in milk of cows fed seaweed can exceed maximum tolerable concentrations, providing Asparagopsis spp. steeped in oil rather than a freeze-dried additive minimises seaweed content, optimises bromoform concentrations, and has resulted in safe milk iodine concentrations (Kinley et al. 2020; Alvarez-Hess et al. 2023; Cowley et al. 2024).
More research is required to determine the impact of bromoform volatilisation from cultivation, processing and animal excretion on ozone, but impacts are expected to be small (Glasson et al. 2022). A full-system bromide balance is also recommended to account for bromoform and its breakdown products, given that bromoform, dibromomethane, bromomethane and bromide are all ozone depleting. This will allow more accurate quantification of true anthropogenic ozone impact to be estimated. Nevertheless, public perceptions may limit seaweed use, especially when even relatively benign chemicals used in agricultural production, such as dicyandiamide (DCD), have been withdrawn from use in Australia when residues were found in milk. A recent publication reported an anthropogenic increase in bromoform in the atmosphere attributed to the rise of power and desalination plants around the world (Jia et al. 2023), suggesting a potential future ban under the Montreal Protocol.
Dietary lipid supplementation is an option currently available to achieve sustained reduction in CH4 production in Australian dairy cattle (Moate et al. 2011). Lipids inhibit rumen methanogenesis directly by reducing methanogen population size, and indirectly by inhibiting rumen protozoal growth, and by decreasing hydrogen in the rumen (Patra 2013). Protozoa produce hydrogen in the rumen and hydrogen is needed for methanogenesis by hydrogenotrophic methanogens (Davison et al. 2020; Nguyen et al. 2020). Methane yield has a negative linear relationship with dietary lipid concentration (Moate et al. 2011). Regardless of lipid type, a 10 g/kg increase in lipid concentration in a typical dairy diet containing ~30 g lipid/kg DMI, reduces enteric CH4 yield by ~3.5% (Moate et al. 2011, 2016).
Lipid intakes greater than 6–7% suppress DMI and, therefore, milk production (Moate et al. 1999; Beauchemin et al. 2008). Therefore, the baseline diet must be low in fat for lipid supplementation to be effective (Williams et al. 2014). Because temperate Australian pastures are low in fat and of poor quality during summer, farmers supplement the diet with feed supplements, hay and silage (Moate et al. 2014; Williams et al. 2014). Because most Australian lipid sources, which include brewers grain, cold-pressed canola oil, cottonseed, grape marc and hominy meal, are by-products, GHG emissions associated with their production can mostly be attributed to the primary product (Williams et al. 2014; Moate et al. 2016). Where basal diets are low in fat, adding lipids to the diet also increases energy intake, and thus, milk production, but not when energy intake remains constant (Grainger et al. 2008, 2010; Moate et al. 2011). The cost-effectiveness of lipid supplementation depends on the cost of lipid by-products relative to the feed replaced and the impact of the lipid by-product-containing diet on milk production compared with the baseline diet (Christie 2019).
Although on-farm emissions reduce during the lipid-supplementation period, emissions reductions are diluted by pre-farm emissions and the proportion of the year in which supplements are fed (Williams et al. 2014; Eckard and Clark 2020). For example, lifecycle analysis has shown that whole cottonseed can be transported 380–1900 km before transport emissions outweigh emissions reductions resulting from supplementation, with a greater distance possible if the fat addition results in a milk production increase (Williams et al. 2014; Ludemann et al. 2016).
Additionally, by-products such as hominy meal or brewers grain are typically fed as part of a total or partial mixed ration (Bramley et al. 2012). Only a small proportion of farms follow a partial or total mixed ration production system and or own mixer wagons required to feed such rations (Wales and Kolver 2017; Murray Dairy 2021). Where farms do not feed mixed rations, starting to feed such rations may intensify the farm system by decreasing the proportion of the diet that comes from grazing pasture and may increase milk production. Therefore, the effectiveness of emissions abatement when introducing lipid by-product supplementation on farms that must purchase equipment and change their farming systems to do so, requires in-depth analysis of any inadvertent emissions changes.
Increasing the proportion of concentrate in the diet has a quadratic relationship with CH4 yield (Moate et al. 2016). Methanogenesis is reduced because an increase in dietary starch relative to neutral detergent fibre shifts digestion from the rumen (where methanogenesis occurs) to the small intestine and increases the amount of propionate relative to the other volatile fatty acids that provide substrates for methanogenesis (Knapp et al. 2014). Comparing concentrate diets, Moate et al. (2017) found that CH4 yields from a single rolled wheat-based diet (10.8 g CH4/kg DMI) were 42%, 60% and 64% lower than those from single rolled corn-based, single rolled barley-based and double rolled barley-based diets respectively, probably because the lower ruminal pH of the wheat diet suppresses ruminal protozoal production of hydrogen, a substrate for methanogenesis. However, in a 16-week trial, Moate et al. (2018) found that a CH4 reduction was not apparent after 10 weeks.
Additionally, as grain-feeding substitutes for pasture consumption (Moate et al. 2020), farmers may increase their cow numbers to improve pasture utilisation. An increase in cow numbers is likely to increase the emissions per unit of land (Christie et al. 2011). Thus, although feeding a higher proportion of concentrates may decrease CH4 yield and emissions intensity, it is likely to increase absolute emissions on Australian dairy farms.
Methane-abating vaccines could be a low-cost, practical and broadly applicable method of reducing enteric CH4, especially in grazing systems (Eckard and Clark 2020; Reisinger et al. 2021; Beauchemin et al. 2022). All stages of the vaccination process have been demonstrated in vitro, but results have not yet been replicated in vivo, despite research starting over 20 years ago (Baca-González et al. 2020; Reisinger et al. 2021; Beauchemin et al. 2022). Once a successful prototype has been developed, a commercial vaccine is likely to take 10 years to reach the market (Reisinger et al. 2021). At this stage, the cost and efficacy of a CH4 vaccine is not known.
Manipulation of calf rumen development could achieve life-long enteric CH4 reductions for a relatively small cost compared with that of ongoing treatment of cows, and would be particularly advantageous in systems where supplementation is not constant (Beauchemin et al. 2022). Early-stage research has shown that a reduction in enteric CH4 persisted for 46 weeks after the cessation of treatments administering 3-NOP to dairy heifers from birth until 14 weeks old (Meale et al. 2021) and for 3 months after the cessation of administering bromochloromethane to goat kids from birth to 3 months old when their mothers were also treated (Abecia et al. 2013). However, Cristobal-Carballo et al. (2021) found that CH4 reductions from feeding chloroform and anthraquinone in the solid feed of dairy calves from 4 days old for 12 weeks were not present 12 weeks after the cessation of feeding. However, Meale et al. (2021) suggested that inclusion of a CH4 inhibitor in solid feed misses the opportunity to influence the rumen microbiome because of inconsistent consumption rates until 3–4 weeks of age. There are still too few studies of early-life programming to understand its efficacy under different conditions (Beauchemin et al. 2022). Future research should aim to minimise methanogen populations in the cows, well prior to parturition, as a key mechanism for establishing low-CH4 progeny is through immune transfer including via colostrum.
Other enteric CH4-mitigating compounds include tannins, saponins, nitrate (NO3−), essential oils, direct-fed microbials and ionophores, such as monensin (Beauchemin et al. 2022). These compounds inhibit methanogenesis directly or indirectly reduce CH4 by modifying the rumen environment (Ku-Vera et al. 2020). More in vivo research is required to identify types, doses, grazing application, productivity impacts, costs and/or risks of these supplements (Cobellis et al. 2016; Moate et al. 2016; Beauchemin et al. 2020, 2022). Dietary NO3− may not be relevant in dairy pasture-based systems where crude protein is already high because of the risk of NO3−-poisoning and DMI decreases (Moate et al. 2016). Trials with monensin have not shown consistent results in mitigating CH4 (Moate et al. 2016). Polyphenolic compounds, tannins, are known to decrease CH4 yield by 0.109 with every g/kg DMI added, but have variable effects on organic-matter digestibility, possibly related to the structural composition of the tannin (Jayanegara et al. 2012; Aboagye and Beauchemin 2019; Beauchemin et al. 2022). Extracts could be used to supplement dairy diets with tannin types that do not affect animal production. Further research is needed to identify specific polyphenolic compounds in key forages and their mode of action on methanogenesis.
Some compounds are available commercially for CH4 abatement with varying results. A feed supplement containing garlic and bitter orange extracts, Mootral (Mootral SA, Rolle, Switzerland), has shown a 10% decrease in CH4 production of dairy cows from a baseline of 575 g/day on total mixed-ration diets without affecting milk yield and 22–26% decreases in CH4 production and CH4 yield of beef steers and Holstein bull calves (Roque et al. 2019b; Brand et al. 2021; Bitsie et al. 2022; Khurana et al. 2023). Meta-analysis showed that essential oil mixture Agolin (Agolin Ruminant, Switzerland), containing coriander oil, geraniol and eugenol, decreased dairy cow CH4 production by 9% (from 227 g/day) and yield by 13% (from 19.7 g CH4/kg DMI) for every gram of supplement fed, and increased fat- and protein-corrected milk yield by 4% when supplemented for more than 4 weeks in total mixed-ration diets (Belanche et al. 2020). Directly fed Lactobacillus spp. had no significant effect on CH4 yield of dairy cows, at doses fed (Williams et al. 2023). Early research has shown that CH4 production reduces by 35% and milk yield increases in dairy cows fed a partial mixed ration including polyphenol-rich sugarcane extract (Polygain™, Poly Gain, Singapore) (Ahmed et al. 2023).
Methane may be reduced through grazing pastures that contain species with specific plant secondary compounds that inhibit CH4, as well as those that improve digestibility or have a high starch or sugar content, as long as the net farm-system emissions are also reduced (Beauchemin et al. 2022; Badgery et al. 2023). In a review of species with potential to reduce enteric CH4, Badgery et al. (2023) highlighted the opportunity to formulate pasture mixes that reduce CH4 and maintain or increase production, and using species grown in Australian dairy systems to abate CH4 would increase the chance of adoption. For example, winter forage rape (Brassica napus L.) fed to lambs had 26% lower CH4 yield over 117 days (from 19.5 g CH4/kg DMI) than did pure perennial ryegrass pasture (Sun et al. 2015). However, where increasing concentrations of forage rape were included in a perennial ryegrass diet, a significant CH4 reduction was not found with a dietary inclusion of <75% (Della Rosa et al. 2024). For a plant species that is found to reduce CH4, it must also meet the agronomic and productivity requirements of a specific grazing system, for adoption (Badgery et al. 2023).
The genetic diversity of some plant secondary compound-containing species already grown in some Australian dairy systems, such as lucerne (Medicago sativa L.), white clover (Trifolium repens L.) and subterranean clover (Trifolium subterraneum L.), may be able to be exploited to deliver low-CH4 cultivars (Kaur et al. 2017; Badgery et al. 2023). Badgery et al. (2023) recommended research focus on improving understanding of plant secondary compounds in commonly used legumes and the ability to enhance them in new cultivars (Sun et al. 2015; Jonker et al. 2017; Della Rosa et al. 2024).
The limited research on CH4-inhibitor and rumen-modifier combinations have shown variable results from different combinations. Substituting lucerne hay for grape marc, a by-product rich in lipids and condensed tannins, Moate et al. (2014) found that grape marc reduced CH4 yield (from 26.1 g CH4/kg DMI) by 23%, ~45% more than expected, than did substituting products containing lipids on their own. Zhang et al. (2021) demonstrated 27%, 32% and 51% decreases in CH4 yield (from 25.9 g CH4/kg DMI) in beef cattle when a baseline diet was supplemented with canola oil, 3-NOP and canola oil with 3-NOP respectively. Likewise, Gruninger et al. (2022) found additive CH4-yield abatement from 3-NOP and oil when fed to beef cattle in combination. They also found distinctly different changes to rumen microbe populations, indicating that 3-NOP and lipids inhibit different methanogenic pathways. Guyader et al. (2015) found synergistic CH4 inhibition from supplementing NO3− and lipids. However, Maigaard et al. (2024) found that the effect of feeding a combination of fat, NO3− and 3-NOP did not result in greater CH4 abatement than did feeding 3-NOP on its own, possibly because of synergistic decreasing effects on DMI. Research is needed to further explore which additive combinations provide additive or synergistic CH4 abatement, and to better understand the interactions among, and approaches to, stacking different additives in grazing-based systems and their economic impacts and feasibility for adoption in different production systems.
Herd management
Genetic herd improvements provide a low-cost, permanent and cumulative method for abating GHG emissions (Richardson et al. 2021a). Abatement can be achieved across a broad scale when a CH4 trait is incorporated into national breeding indices (Rowe et al. 2021). Phenotypic variation has been identified in dairy cow CH4 traits for CH4 production, CH4 yield, CH4 intensity and residual CH4 production (de Haas et al. 2017). Residual CH4 production, which identifies differences in CH4 production in cows with similar milk production and bodyweight, is the easiest to include in a breeding index without affecting other economically important traits (Manzanilla-Pech et al. 2021). Methane production can also be abated by changing the emphasis of currently included traits, such as survival traits, although at much lower rates of abatement than inclusion of the CH4 trait (Richardson et al. 2022).
A sustainability index was established in 2022 for Australian dairy farmers to select bulls to lower their herd emissions intensity by changing the emphasis of traits already included in current Australian breeding indices (Nguyen et al. 2023). Expected emission-intensity (CO2e/kg protein equivalent) abatement from the sustainability index was 5.5–7.6% by 2050, relative to 2015 levels, whereas expected emissions per cow were not reported (Nguyen et al. 2023). Meanwhile, modelling the incorporation of an accurate CH4 trait into an Australian multi-trait dairy index that allows for continued genetic gain in other areas, Richardson et al. (2022) found that emissions intensity (CO2e/kg protein equivalent) was reduced 21.25% and emissions per cow were reduced 8.23% from 4298 kg CO2e over a 30-year breeding program, when the economic value placed on carbon within the index was AU$250/t. The differences in emission-reduction potential highlight the need for a CH4 trait to be developed for incorporation into breeding indices.
The main challenge for development of a CH4 trait is that CH4 production must be quantified on thousands of animals on commercial farms. Either technology to measure CH4 on large numbers of animals is required or proxies for CH4 production must be developed (de Haas et al. 2017; Manzanilla-Pech et al. 2021). Although research is underway to develop appropriate proxies, a CH4 trait that can be incorporated into commercial breeding indices is not yet available (de Haas et al. 2017; Richardson et al. 2021b).
Changes to breeding programs must also consider unintended consequences. For example, breeding for emission-intensity abatement can result in increases in absolute emissions (González-Recio et al. 2020; de Haas et al. 2021). Meanwhile, breeding for absolute emissions abatement alone can result in smaller cows that produce less milk, whereas breeding for emissions intensity results in larger cows with a greater milk-production potential (López-Paredes et al. 2018; Manzanilla-Pech et al. 2021). Therefore, selection of CH4 traits for potential incorporation into dairy breeding indices must take account of other economically important traits and impacts on different CH4 metrics (Manzanilla-Pech et al. 2021). Additionally, the effects of breeding for CH4 abatement on diets and animal health need to be further understood (Beauchemin et al. 2020; Manzanilla-Pech et al. 2021).
In typical south-eastern Australian dairy herds, cows calve once a year, with 10 months of lactating in either a seasonal or split-calving system, based on seasonal pasture growth (Malcolm 2005; Dairy Australia and Agriculture Victoria 2021; Datagene 2021). In an extended lactation system, cows lactate for longer than 10 months (Auldist et al. 2007; Browne et al. 2014). Auldist et al. (2007) found that cows that lactated for 16 months produced 99% of the annualised milk solids of those lactating for 10 months, and the same number of cows reached the target lactation length in each case. However, a lactation length greater than 16 months resulted in fewer cows meeting the target lactation length.
When modelling the effect of a 16-month compared with a 10-month lactation on cow lifetime GHG emissions, Browne et al. (2011) found that the extended lactation resulted in lower absolute emissions and emissions intensity, mostly attributed to a 9% decrease in the heifer-replacement rate. However, keeping the heifer-replacement rate constant, Christie (2019) found that extended lactation resulted in increased absolute emissions, but decreased emissions intensity because annual milk production was greater under the extended lactation system than in the baseline farm. Therefore, CH4 reductions from extended lactation systems are a function of reducing the heifer-replacement rate, which can also slow genetic gains, decrease opportunities to cull older and more emission-intensive cows, and shift GHGs off-farm as excess young stock are sold to other farmers (Knapp et al. 2014; Christie 2019).
Paddock-derived nitrogenous emissions
To produce high pasture dry matter (DM) yields requires significant N inputs (Eckard and Franks 1998; Rawnsley et al. 2019). This N can be from external sources (fertiliser, legumes, imported feeds) or internally recycled sources (dung, urine, mineralisation). Dairy pastures in Australia receive, on average, 214 kg N/ha.year from external sources (Stott and Gourley 2016), with more heavily stocked farms receiving as much as 426–506 kg N/ha.year from all N sources (Eckard et al. 2007; Rawnsley et al. 2019). These N-fertiliser rates are similar in New Zealand, with an average of 156 kg N/ha being applied on System 5 farms and average N inputs totalling 274 kg N/ha in Canterbury (Pinxterhuis and Edwards 2018; Serra and Pinxterhuis 2020). This compares with other high-N fertiliser systems such as cotton (180–240 kg N/ha.year, Rochester and Bange 2016) and sugar cane (164–240 kg N/ha.year, Thorburn et al. 2017).
However, fertiliser N inputs contribute less than half of the total N inputs into a pasture-based dairy system, with dung and urine being the main inputs of N (Stott and Gourley 2016). About 70–90% of N ingested by dairy cows is excreted, mainly as urine (Haynes and Williams 1993). Of this excreted N, more than 60% is not recycled and is therefore lost to the environment (de Klein and Eckard 2008). Eckard et al. (2007) showed that the N surplus on a dairy farm could increase from 79 kg N/ha.year, with no N fertiliser added, to 218 kg N/ha.year with 200 kg N/ha.year applied. Stott and Gourley (2016) reported a consistent increase in the N surplus across the Australian dairy industry from 54 kg N/ha.year in 1990 to 158 kg N/ha.year in 2012, with similar N surplus of 180 kg N/ha.year reported from the DairyBase analysis in New Zealand (Pinxterhuis and Edwards 2018).
Simply adding more N inputs to achieve higher pasture growth is no longer an acceptable solution, given the exponential relationship modelled between N input and N loss (Eckard et al. 2006; Christie et al. 2020). The regulation of N surpluses in New Zealand, Ireland and the Netherlands (see New Zealand Government (2021), TEAGASC (2023) and Kros et al. (2024)) should act as a warning to the Australian dairy industry. Meanwhile, the contribution of N loss to Australia’s dairy carbon footprint is relatively high compared with many countries globally because of our reliance on grazed pastures (Mazzetto et al. 2022). Reducing the N surplus (N inputs minus outputs in product) in dairy production systems is therefore a pressing research priority in the next decade.
Nitrogen-loss pathways
Environmentally significant N losses are caused by nitrification and denitrification processes in the soil that result in the formation of N2O, NO3− or ammonification. Whereas N2O contributes directly to GHG accumulation, NO3 leaching and ammonia (NH3) volatilisation also contribute indirectly to N2O emissions when re-deposited offsite (de Klein and Eckard 2008). Additionally, NO3− is a water pollutant, whereas NH3 volatilisation is not considered environmentally significant in Australia (Eckard et al. 2007). External inputs that particularly contribute to these processes are nitrogenous fertilisers and animal excreta, with urine having a much larger impact than dung (de Klein and Eckard 2008). The rate and extent of direct and indirect N2O production is influenced by soil water content, temperature and soil pH, carbon and NO3− (Whitehead 1995; de Klein and Eckard 2008). Therefore, N2O abatement is particularly dependent on changes in the fertiliser or urine inputs and creating conditions that prevent N2O production.
Dairy pastures produce 1.3–7.2 kg N2O-N/ha.year (0.4–2.0 t CO2e/ha.year; Eckard et al. 2003; Phillips et al. 2007). Using data from the Australian National GHG Inventory database (Australian Government 2022), combined with Australian Bureau of Statistics land-use data (ABS 2022), dairy farming would be the largest emitter of N2O per hectare in Australia, estimated at 2.46 t CO2e/ha.year, compared with other high N-input industries such as sugar cane at 1.97 t CO2e/ha.year and cotton at 1.33 t CO2e/ha.year. de Klein and Ledgard (2005) reported similar emissions of 2.1–2.6 t CO2e as N2O from dairy farms in New Zealand.
Australian dairy farms are either irrigated or located in higher-rainfall regions; this, coupled with high stocking rates, makes dairy pasture soils vulnerable to drainage and therefore leaching of NO3− (Eckard et al. 2004, 2007). Annual NO3− leaching loads, measured from dairy pastures on poorly drained soils in Australia, were 4–15 kg and 22–38 kg NO3−N/ha, from the application of 0 and 200 kg N/ha respectively (Eckard et al. 2004; Dougherty et al. 2016). Although these NO3− loads are lower than those reported from higher-rainfall regions on more free-draining soils in New Zealand (Burkitt 2014), these data are similar to NO3− leached from sugar cane, on sandy soils in high-rainfall coastal regions of northern Australia (Rasiah et al. 2003; Armour et al. 2013). This adds a further 1–10 t CO2e/ha.year to the GHG footprint of pasture-based dairy production.
Although N leaching on Australian dairy farms is technically no different from that on New Zealand farms, for a given rainfall and soil-drainage capacity, the environmental impact in New Zealand is compounded by the lack of diverse land use within most dairy catchments (Rawnsley et al. 2019).
Apart from water, urine from ruminants largely comprises N, predominantly in the form of urea (Whitehead 1995). Most dairy pastures and soils would have liberal amounts of urease enzyme present, leading to potential for NH3 volatilisation from both urine or urea fertiliser (Eckard et al. 2004; Smith et al. 2020). Eckard et al. (2003) reported NH3 volatilisation from a grazed dairy pasture at 17–57 kg NH3-N/ha.year, where pastures were fertilised with 0 and 200 kg urea N/ha.year, and stocked at 1.9 and 2.8 cows/ha respectively. Dairy pastures therefore would lose more NH3 (from urea fertiliser plus urine) through volatilisation than would any broadacre agricultural industry in Australia (apart from confinement livestock systems such as feedlots). Ammonia volatilisation adds another 0.3–1.0 t CO2e/ha.year to the GHG footprint of dairy farming. Given the magnitude of N losses through NH3 volatilisation, there is both a financial and environmental imperative to provide dairy farmers with options to reduce NH3 volatilisation.
Fertiliser- and soil-derived N-loss abatement
Current fertiliser practice on dairy farms is to apply N fertiliser at a set rate of between 30 and 50 kg N/ha after each grazing (Smith et al. 2018). However, Christie et al. (2018b) found that the N-fertiliser rate required to produce 90% of relative pasture yield varied according to location and season. By lowering N-fertiliser rates to achieve 90% relative pasture production compared with 100%, Christie et al. (2020) found that N fertiliser losses could be reduced by 34–74%, depending on season or location. Additionally, comparing set-rate strategies with strategies that responded to soil or plant concentrations, Smith et al. (2018) found minimal production impacts and 2–75% lower N losses with ~20–50% annual N-fertiliser reduction, depending on fertiliser strategy and location. More research into plant N thresholds that determine N timing and rate requirements is needed for such strategies to also optimise production.
Variable rate applications aim to reduce N-fertiliser amount by targeting fertiliser placement according to demand. Variable-rate fertiliser applications using technology that avoids spreading fertiliser on urine patches has been shown to reduce N-fertiliser input by 10–50% (Hills et al. 2014; Rawnsley et al. 2019). Given the economic cost of N fertiliser to farmers and the need to reduce N emissions, further research and extension of methods that reduce N fertiliser in the absence of production declines should be a research priority.
Enhanced-efficiency fertilisers slow or inhibit soil N transformations (Chen et al. 2008). They include nitrification and urease inhibitors, which prevent ammonium conversion to NO3− and slow urea hydrolysis to ammonium respectively (Di and Cameron 2016; Li et al. 2018). Decreasing soil NO3− prevents NO3− leaching and N2O production, and decreasing NH3 volatilisation prevents NH3 deposition and indirect N2O emissions (de Klein and Eckard 2008). Commercially available products include the nitrification inhibitors DCD and 3,4-dimethylpyrazole phosphate (DMPP), and the urease inhibitor N-(n-butyl) thiophosphoric triamide (NBTPT).
Urease inhibitors result in a 20–88% decrease in ammonium volatilisation (Suter et al. 2016). Adoption of nitrification inhibitor-coated fertiliser could reduce N2O from N fertiliser by as much as 80% (de Klein et al. 2010; Suter et al. 2016). However, the N2O abatement has not always translated to commercial pastures (Nauer et al. 2018). Effectiveness of either inhibitor is dependent on formulation, climate, soil type, land use and baseline emissions (Chen et al. 2008; Lam et al. 2017).
Additionally, a decrease in one N-loss pathway can result in an increase in another (Rawnsley et al. 2019). Ammonia retention through urease inhibitor use can either decrease or increase N2O emissions (Suter et al. 2016; Li et al. 2018), and nitrification inhibitors can result in higher NH3 volatilisation that can sometimes result in a net increase in N2O emissions when accounting for indirect emissions (Lam et al. 2017). Both inhibitors may need to be used in combination to reduce direct and indirect N2O emissions (Lam et al. 2017). Meta-analysis has shown that the use of both inhibitor types together decreased direct and indirect N2O losses by 50%, on average (Li et al. 2018).
Enhanced-efficiency fertiliser costs more than conventional urea fertiliser and is unlikely to deliver a productivity gain (Eckard and Clark 2020). However, using enhanced-efficiency fertilisers in combination with decreasing N-fertiliser rates may be a cost-effective way to best utilise N fertilisers (Li et al. 2018; Nauer et al. 2018; Rose et al. 2018; Rawnsley et al. 2019). Productivity was maintained in subtropical dairy pastures where DMPP was combined with half the conventional urea rate compared with the conventional urea rate alone (Rowlings et al. 2016). On southern Australian dairy pastures, productivity was maintained during spring when DMPP was combined with 34% of the conventional urea rate (Suter et al. 2022). Given that DMPP-amended urea costs ~20–30% more than does conventional urea, Nauer et al. (2018) estimated that farmers could save 30% of fertiliser costs by applying DMPP-amended urea at half the conventional rate. In the future, all N fertilisers may be coated with an inhibitor. For these fertilisers to be cost-effective, research should focus on the extent of fertiliser-rate reductions possible, while maintaining productivity and the effects of application on all N-loss pathways in different conditions.
Urine-derived N-loss abatement
Mitigating N2O losses from urine deposition is a more challenging research priority, given the high concentrations of urine deposited on limited areas of the pasture at any point in time. The effective N application rate within a urine patch from a dairy cow grossly exceeds the capacity of the soil–plant system to effectively utilise the N deposited. Methods to abate direct and indirect N2O emissions from urine patches focus on direct treatment of patches or changes in feed to influence N output or interactions between N in urine and other urine components.
The crude protein requirement of a dairy cow during lactation ranges from 12–15% to 16–18% in late versus early lactation respectively (Moran 2005). The crude protein in well-fertilised dairy pastures can be greater than 20%; so, matching cow N requirements with feed N supply can be difficult in a pasture-based system (Powell et al. 2010; Chapman et al. 2012; Lucas 2015; Cullen et al. 2017). Low-N feed supplements can reduce wasted N and N2O emissions (Misselbrook et al. 2005; Christie et al. 2014; Lucas 2015). Given that pasture is supplemented with grain in Australian systems, Australian dairy has a natural advantage over New Zealand grass-based dairy systems to adjust the dietary energy to protein ratio. Additionally, the increasing use of feedpads, partial mixed rations and confinement feeding present options for more intensive dairy systems to recycle excreted N back to pasture or utilise low-N high-energy forages, for example, maize silage.
By reducing the crude protein in a dairy cow diet from 19% to 14%, Misselbrook et al. (2005) found a 47% decrease in urinary N (from 8.5 to 4.5% g/L). Modelling dairy cow diets in a temperate medium-rainfall location, Christie et al. (2014) predicted a 52% decrease in direct and indirect N2O emissions (from up to 4.4 t CO2-e/ha.year) as crude protein was decreased from 26% to 15%, and Smith et al. (2021) found that the overall N volatilisation from all N sources could be reduced from an average of 52 kg N/ha.year by 47%, by balancing the dietary energy to protein ratio.
However, N-use efficiency can decline when milk production is compromised by low-protein feeds that are also low in energy, so that dietary energy levels must be maintained (Christie et al. 2014). Modelling the effect of reducing dietary N on GHG emissions on four case study farms by replacing half the baseline grain in the diet with maize silage, Christie (2019) found that results were dependent on the baseline diet. Emissions increased where the baseline diet already had a high DM digestibility to protein ratio, with DM digestibility reductions resulting in reduced milk production.
To inhibit N losses from urine patches, research trials have either sprayed nitrification inhibitors onto urine patches or orally fed nitrification inhibitors to the dairy herd. International assessments have shown that a nitrification-inhibitor urine treatment abates direct N2O by 51–62%, and up to 82% in some individual studies, and NO3− leaching by 30–50% (Di and Cameron 2002, 2016; Soares et al. 2023). Australian studies have shown that N2O abatement from urine patches has been lower than in international studies. On northern Victorian irrigated perennial ryegrass (Lolium perenne L.)-based pastures, Kelly et al. (2008) found that DCD sprayed onto urine patches reduced urine-patch N2O emissions by 47% in spring (from 5.6 kg N2O-N/ha) and 27% (from 4.7 kg N2O-N/ha) in summer. On western Victorian perennial ryegrass pastures, Ward et al. (2018) found that spraying nitrification inhibitor, nitrapyrin, onto urine patches decreased N2O emissions by 0–35% (from 5.35 kg N2O-N/ha). Nitrapyrin was effective at decreasing N2O emissions when soil water content was below 40% or above 90%.
Although nitrification-inhibitor sprays are effective at reducing N losses from urine patches, the practicalities of doing so must be addressed. A recently developed New Zealand technology, Spikey® (Pastoral Robotics, New Zealand) that identifies and sprays urine patches with N inhibitors while applying N fertiliser to the remaining paddock area, must be applied within 2 days (Quin 2018; Pastoral Robotics 2023). Use of such technology requires a large initial investment in equipment or ongoing lease cost and ongoing nitrification inhibitor spray costs and will have limitations where paddocks are too wet for a tractor, or when cows graze paddocks for longer than 2 days. Although such technology is an example of the kind of practical solution required for nitrification-inhibitor sprays to be used on farms, in the absence of subsidies, there is currently no economic reason for Australian farmers to use nitrification-inhibitor sprays.
Oral administration to dairy cows is another method of delivering nitrification inhibitors directly to urine patches. Oral administration of DCD has reduced N2O emissions from urine patches to the same extent as has mixing DCD with urine (Luo et al. 2015). After feeding New Zealand cows 30 g DCD per year, Luo et al. (2015) found that dairy cow urine patches contained 60 kg/ha DCD and reduced autumn N2O emissions by 40%. This had the same result on N2O emissions as had treating urine patches with 60 kg/ha DCD (Luo et al. 2015). After mixing DCD into water troughs so that DCD intake was 46–110 g DCD per animal per day, Welten et al. (2014) found that New Zealand dairy cow urine-patch N2O emissions were reduced by 45%. However, the DCD feeding rates tested by Luo et al. (2015) and Welten et al. (2014) are orders of magnitude higher than when DCD-coated fertiliser is applied to pasture (Ray et al. 2023). Therefore, use is unlikely until concerns about residues have been resolved and there are guarantees that trade will not be affected.
Several other mechanisms may reduce N loss from urine. Feeds with high water content can increase urination events, spreading out N deposition so that each urination event contains less N (Bryant et al. 2020). Dung from dairy cattle has a lower emissions factor than does urine; so, increasing N partitioning to dung rather than urine reduces N2O emissions (de Klein et al. 2020; van der Weerden et al. 2020). Additionally, secondary metabolites contained in plant species can act as nitrification inhibitors (de Klein et al. 2020). Plantain pastures have been shown to reduce N losses compared with perennial ryegrass pastures, most convincingly through their higher water content (Minnée et al. 2020; Eady et al. 2024; Fransen et al. 2024). Some studies have reported that plantain content in ryegrass pastures has increased the proportion of N partitioned to dung (Minnée et al. 2020) and may contain nitrification-inhibiting secondary compounds (Fransen et al. 2024). Other studies have reported pastures containing at least 30% plantain to maintain or increase milk production with and lower N2O emissions by at least 30% and up to 90% compared with perennial ryegrass, lucerne and white clover pastures (Box et al. 2017; Luo et al. 2018; Minnée et al. 2020; Rodríguez Gelós 2020; Vi et al. 2023). Nitrate leaching was reported to reduce 20–60% when plantain comprised 30–40% ryegrass–white clover pasture DM (Pinxterhuis et al. 2024). However, research on the nitrogen mitigation effects of plantain are currently the subject of debate (Eady et al. 2024; Fransen et al. 2024). Further corroborative research is required in other dairy systems before the debate can be resolved.
Using pasture species to abate nitrogenous emissions is potentially a low-cost option without major changes to current practices (Fransen et al. 2024). Plantain is one of a number of pasture species available to fill the summer and autumn feed gap in Australian pasture-based dairy systems (Özkan et al. 2015). However, where plantain is planted, it is commonly only on a proportion of the farm, so the potential emissions reduction would be diluted when extrapolated out to a GHG reduction per hectare annually. Additionally, issues around persistence, weeds and pest control and palatability must be addressed for increased adoption on commercial farms (Pinxterhuis et al. 2024). Identification of pastures that can mitigate nitrogenous emissions and fit within Australian pasture systems should be an area of future research.
As well as contributing to the enteric CH4 abatement, plant secondary metabolites, such as hydrolysable and condensed tannins, result in increasing the efficiency of dietary-N absorption, increasing partitioning of N excretion from urine to dung, which has a lower emissions factor and increases nitrification-inhibitor excretion (Mueller-Harvey 2006; Grainger et al. 2009; Gao and Zhao 2022). More in vivo research is required, as well as research into toxic and antinutritional effects, identification of combinations of secondary compounds that reduce N losses, that are palatable and do not reduce dietary intake, and their effects on direct and indirect N2O emissions (Reisinger et al. 2017; Gao and Zhao 2022; Kreuzer 2023).
Salt supplementation increases water intake, urinary volume and number of urine patches, and decreases the urinary N concentration (Ledgard et al. 2007, 2015; Dijkstra et al. 2013). Ledgard et al. (2015) found that increased urine events and decreased urinary N concentration resulting from salt supplementation could decrease N leaching by 10–22% across a paddock. However, high-salt diets can decrease milk production and increase salt concentrations in soil and effluent (Spek et al. 2012; Ledgard et al. 2015; Ben Meir et al. 2023). Research would be required on the long-term health effects and public acceptance of feeding salt.
Breeding for low urinary N excretion or improved N-use efficiency have been suggested as methods to reduce urinary N loads (Beatson et al. 2019; Aizimu et al. 2021; van den Berg et al. 2021). Although both traits are difficult to measure on large numbers of animals, milk urea N has been suggested as a proxy (Richardson et al. 2023; Tavernier et al. 2023). Milk urea N can be easily measured on large numbers of animals. However, correlation between milk urea N and urinary N excretion or N-use efficiency have varied (Aizimu et al. 2021; Handcock et al. 2021; Tavernier et al. 2023) and heritability is low in Australian dairy cattle (van den Berg et al. 2021). Additionally, although Ariyarathne et al. (2021) found that selecting for reduced milk urea N using the New Zealand national dairy breeding index resulted in a lower milk urea N per cow, stocking rates increased compared with no milk urea-N selection, resulting in no difference in N excretion per hectare.
On-off grazing is a tool used during winter and early spring in wet regions of south-eastern Australia to decrease compaction and waterlogging damage from prolonged pasture grazing (Sinnett et al. 2023). The milking herd grazes for 2–6 h per day and is then moved to a stand-off area such as a feed pad (Sinnett et al. 2023). Reducing the time non-lactating dairy cow grazed from 24 to 6 h per day during winter and early spring reduced N2O emissions by 40–65% (baseline 0.225 and 0.968 kg N2O-N/ha; Luo et al. 2013). In lactating New Zealand dairy cows during autumn, reducing grazing to 3 h per day reduced both direct N2O emissions and NO3−-leaching by about 40% and direct and indirect on-farm annual N2O emissions by 7–11% (baseline 1.4 kg N2O/ha; de Klein et al. 2006). The New Zealand studies involved on-off grazing during wet periods (de Klein et al. 2006; Luo et al. 2013); so, the effectiveness on N2O reductions will depend on how wet the winter is expected to be. Reisinger et al. (2018) concluded that on-off grazing would reduce N2O emissions only on farms with poorly drained soils with more than 150 wet days per year.
Although direct N2O and N leaching on Australian dairy farms is technically no different from that on New Zealand farms, for a given rainfall and soil-drainage capacity, the environmental impact in New Zealand is compounded by the lack of diverse land use within most dairy catchments (Rawnsley et al. 2019). Unless research can deliver solutions to reduce the N loading on dairy soils, the options would be either reducing stocking rate or increasing the diversity of land use within dairy catchments, to effectively dilute the N2O and NO3− leached before entering streams or groundwater. Both options have significant financial implications for dairy farmers, the former in reduced milk production from fewer cows, and the latter implying the purchase of, or payment for, sink-land in dairy catchments.
Effluent management-derived emissions
In grazing-based systems, effluent from the milking shed, adjacent concreted yards and, sometimes, a feeding area, is washed into effluent ponds (Houlbrooke et al. 2004; Gourley et al. 2012; Laubach et al. 2015). In some systems, effluent is washed into a trafficable solids trap, with the liquid fraction entering the effluent pond and the solid fraction being periodically removed and stockpiled for spreading (Laubach et al. 2015). The liquid fraction is generally stored in two storage ponds adjacent to the milking shed (Gourley et al. 2012). Liquid effluent is sprayed onto pastures after storage in ponds (Laubach et al. 2015). Methane, N2O and NH3 are all lost at various stages of manure-management processes (Gerber et al. 2013). The main GHG produced from manure storage is the CH4 produced from effluent ponds (Laubach et al. 2015). Methane is also produced from solids-separator manure, whereas little is produced once manure is applied on land (Montes et al. 2013; Laubach et al. 2015). Although making smaller contribution to total CO2e from manure management than pond CH4 production, direct N2O emissions can be produced by stored solids and land application of effluent, and indirect N2O emissions can be produced through NH3 volatilisation from yards, solids storage, effluent ponds or effluent land applications once NH3 re-deposits elsewhere (Laubach et al. 2015). Abatement of one GHG can induce production of another; so, abatement strategies must account for effects on all pollutants if they are to result in a reduction in net emissions (Schils et al. 2013).
Compared with countries where housed dairy systems dominate, Australia’s mean proportion of effluent-derived emissions is small (Mazzetto et al. 2022). Additionally, a lack of economic incentive to abate effluent emissions makes strategies to reduce manure emissions cost prohibitive for most Australian dairies (Dairy Australia 2022). Therefore, emissions reduction from effluent management should be a lower research priority than that from other emissions sources.
Covered anaerobic ponds
Covering effluent ponds with an impermeable cover creates an anaerobic environment, initiating anaerobic digestion and producing biogas, a mixture of CH4 and CO2 (Montes et al. 2013; Heubeck 2015). Methane from covered anaerobic ponds can be flared to abate emissions or used for energy production and anaerobic digestion removes carbon, indirectly reducing N2O emissions from effluent application to pasture (Heubeck 2015; Laubach et al. 2015).
The amount of CH4 available for capture depends on the amount of volatile solids reaching the pond, which ranges from ~15% in pasture-based systems to 94% in a total mixed-ration system (Grell et al. 2024). Effluent systems with passive solid–liquid separation, where trafficable solids traps are not cleaned out, can reduce volatile solids entering effluent ponds further because CH4 is lost in anaerobic conditions created when manure is held as slurry (Grell et al. 2024). A covered anaerobic pond can remove ~86% of the volatile solids added (Heubeck and Craggs 2010). Industry analysis has shown that the CH4 abated from covered anaerobic ponds would reduce annual CO2e production by ~5% or more, depending on the system intensity, but was not likely to be economically feasible for farms with fewer than 1000 cows in a total mixed-ration system (Dairy Australia 2022).
Flocculation
Treating effluent that leaves the dairy shed, yard or feedpad to remove organic matter can reduce fugitive CH4 emissions from anaerobic effluent storage (Amon et al. 2006). Grell et al. (2023) found that effluent separated with a Z-filter solid–liquid separator and treated with lime and cationic polymer flocculant could remove 85% of volatile solids (from a baseline of 0.45% wet), while increasing the fertiliser value of separated manure. The impact of treatment on other GHGs was not assessed. Treating effluent with polyferric sulfate, a coagulant used to flocculate colloidal particles in dairy effluent and produce clarified water, reduced CH4 by 99% (from 1051 g CO2e/m2), with small changes to other GHGs in the effluent and water fractions over a 2-month treatment period (Cameron and Di 2021; Chisholm et al. 2021). Polyferric sulfate treatment may result in inhibition of methanogens, sulfate- and ferric-reducing bacteria outcompeting methanogens for substrate, and anaerobic oxidation of CH4 (Cameron and Di 2021). However, systems would require large capital investment, limiting possible adoption, and more research is needed to understand the mechanisms behind the abatement and the impact in different effluent treatment systems.
Nitrification inhibitors
Mixing a nitrification inhibitor with dairy effluent may reduce N2O emissions from dairy effluent by slowing the rate that ammonium converts to NO3− (Li et al. 2014). However, studies in New Zealand have shown mixed results. After mixing nitrification inhibitor, DCD, with fresh and stored effluent and fresh manure just prior to application, Li et al. (2014) found N2O emission reductions of 46–90%. Comparing the effect of DCD mixed with fresh dairy effluent, fresh manure, stored dairy effluent and stored manure, Li et al. (2015) observed N2O reductions of 24–84%, depending on the season. Both of these studies were located at the same single site. However, across four New Zealand sites, van der Weerden et al. (2016) found no difference between dairy effluent with and without DCD. A better understanding of nitrification inhibitor effects on dairy effluent is therefore needed to establish whether it is an effective emission-abatement method.
Acidification
Effluent acidification has been shown to reduce CH4 emissions and NH3 volatilisation from stored manure, and to reduce N2O emissions and NH3 volatilisation after application to pastures and may improve fertiliser N value (Fangueiro et al. 2010, 2013, 2015). Methane emissions from slurry storage have been reduced by 17% to over 90%, depending on the acid used and may improve fertiliser N value (Fangueiro et al. 2015). Acidification is used commercially in Denmark on ~20% of farms (Walsh 2022). Commercial manure-acidification systems mix acid with slurry held in slurry-storage tanks or add acid to slurry spreaders during application (Fangueiro et al. 2015). Effluent acidification would need to be tested on Australasian effluent-treatment systems and pasture-management systems to quantify effects on emissions during storage and after application to pasture. Potential commercialisation would also require an economic incentive for farmers.
Sequestration
Tree planting
Doran-Browne et al. (2018) showed that trees could play a significant role in offsetting emissions on south-eastern Australian sheep and beef farms over a 25-year period. However, dairy farms occupy some of the best rainfall and soil-type combinations for productive agriculture and are hence some of the most cleared in Australia (Eckard and Clark 2020; Adams and Engert 2023). Thus, the high value of agricultural land prevents large areas of tree establishment as a carbon sink on dairy farms (Reisinger et al. 2017). Nevertheless, trees may be planted in paddocks, in riparian zones, along paddock edges and in small non-productive areas, without affecting production (Reisinger et al. 2018). Shade provided by shelter belts and paddock trees reduces cow cold and heat stress and may improve milk production (Kendall et al. 2006; Fisher et al. 2008; England et al. 2020). Whereas small areas of trees provide benefits to production and offset emissions, larger emissions reduction from tree plantations would require securing lower-value land, deliberately for tree planting. Further analysis is required to understand the configuration of trees to maximise co–benefits versus just carbon sequestration alone.
Soil carbon
There is local and international interest in increasing soil carbon on agricultural soils to abate GHGs (Abbas et al. 2020; White 2022). However, the potential of agricultural soils to store carbon depends on how far the soil is from reaching an equilibrium state (Wang et al. 2023). The high-rainfall and productive soil-type environment where most dairy farms are located, coupled with permanent pastures, means that the soil carbon on Australian dairy farms is high and relatively stable (Schipper and Sparling 2011; Eckard and Clark 2020). Additionally, management practices that improve soil carbon stocks under intense grazing systems have not been clearly identified and practices that may increase soil carbon may decrease it elsewhere or compromise milk yield (Kirschbaum et al. 2017; Whitehead et al. 2018).
Energy-derived and pre-farm embedded emissions
Options to reduce energy-derived and pre-farm embedded emissions are increasingly available and the unit costs are reducing, improving the economic incentive for adoption. Australian and New Zealand reports show ~20% electricity reductions possible on dairy farms from energy-equipment upgrades (Lucas 2015; Ilyas et al. 2020). Solar energy is already installed on ~50% of Australian dairy farms, although there are limitations to its use without battery-cost reductions because energy generation periods do not match major demand periods (Best and Burke 2023; Dairy Australia 2023b).
Future emissions reductions may come through fossil-fuel reductions on farm or through fertiliser-embedded emissions. For example, N fertiliser produced by clean energy, green NH3, has potential to eliminate pre-farm emissions from N-fertiliser production (The Royal Society 2020; Ojelade et al. 2023). Hybrid, battery electric and fuel-cell electric tractors are also in the initial stages of development once current runtime and horsepower challenges are overcome (Lagnelöv et al. 2020; Stakens et al. 2023).
Discussion and conclusions
Although the pasture-based dairy industries in Australia and New Zealand have significant community support, this social licence is tentative and increasingly vulnerable, given the environmental impact under a changing climate. Scope 3 targets set by Australian and international dairy processors and retailers make it imperative that research deliver options so that farmers can substantially reduce their environmental footprint (largely CH4 and N losses). Enforcement of Scope 3 targets would rely on farmers implementing options to abate their emissions in line with other exporters in competing markets, to ensure a future for Australian dairy exports in supply chains that prioritise sustainability. Pressure to reduce emissions will also come with new legislation. For example, new Australian Government climate-related financial disclosure legislation will introduce mandatory GHG emission reporting, including Scope 3 emissions, for large companies from January 2025 (Parliament of Australia 2024; Zhou 2024). Although Australian trading partners do not have current plans to include agricultural goods in potential border-adjustment mechanisms, there are large risks to Australian agriculture if such policies are implemented (Deloitte Access Economics 2023), further highlighting the need for Australian dairy to be prepared for market-driven emission-reduction requirements.
This paper has provided an overview of potential GHG abatement strategies relevant to Australian dairy farms. Table 1 provides a summary of the estimated abatement potential of the different strategies discussed that can be expected to be available in the near term, that abate absolute emissions, and either have no or a positive impact on milk production and are feasible options on Australian dairy farms. Some strategies have been excluded from the summary because they have been shown not to cause GHG abatement (e.g. extended lactations), have not yielded consistent results (e.g. soil carbon or nitrification inhibitors in effluent systems), and have animal-health implications unable to be overcome (e.g. salt supplementation) or implications for cow numbers that would result in increased absolute emissions (e.g. concentrate feeding).
| Strategy | Absolute GHG abatement (%)A | Certainty | Availability | Key challenges | Key references | |||
|---|---|---|---|---|---|---|---|---|
| Enteric CH4 | Direct and indirect N2O | Waste CH4 | ||||||
| 3-NOP | 30–40 | – | – | Medium | Available | Grazing system applicability; cost | Dijkstra et al. (2018), Kebreab et al. (2023) and Costigan et al. (2024) | |
| Seaweed | 40 | – | – | Medium | Available | Consumer acceptance; cost | Lean et al. (2021) and Sofyan et al. (2022) | |
| Lipid by-products | 7B | – | – | High | Available | Life-cycle emissions effected by transport distance and time supplemented; system infrastructure | Williams et al. (2014) and Moate et al. (2016) | |
| Early-life programming | 17.5 | – | Low | Unknown | Replicability | Meale et al. (2021) | ||
| Garlic and bitter orange extract | 10–25 | – | – | Low | Available | Grazing system applicability; cost | Roque et al. (2019b), Brand et al. (2021), Bitsie et al. (2022) and Khurana et al. (2023) | |
| Essential oil mix containing coriander oil, geraniol and eugenol | 9 | – | – | Medium | Available | Grazing system applicability | Belanche et al. (2020) | |
| Polyphenol-rich sugarcane extract | 35 | – | – | Low | Available | Number of studies | Ahmed et al. (2023) | |
| Breeding index with an accurate CH4 trait | 0.2–0.4 CO2e per year, cumulative | – | – | Medium | Unknown | CH4 measurement; interactions with other economic traits, diet and animal health | Manzanilla-Pech et al. (2021) and Richardson et al. (2022) | |
| Decreased N-fertiliser application | – | Decreased N losses: 34–74 | – | High | Available | Extension; applicability on farms | Christie et al. (2020) and Smith et al. (2018) | |
| Variable-rate fertiliser application | – | Decreased N fertiliser application: 10–50 | – | Low | Available | Capital cost; applicability on farms | Hills et al. (2014) and Rawnsley et al. (2019) | |
| Urease inhibitors | – | Decreased NH3: 20–88 | – | High | Available | Cost, lack of productivity increase | Suter et al. (2016) and Li et al. (2018) | |
| Nitrification inhibitor-treated fertiliser | – | Direct N2O: 8–80 NO3− leaching: 35–52 NH3: −3 to −65 | – | High | Available | Cost; quantification of fertiliser reduction possible; regulation of DCD; lack of productivity increase | de Klein and Eckard (2008), Suter et al. (2016), Lam et al. (2017) and Li et al. (2018) | |
| Dietary energy:protein ratio | – | Direct and indirect N2O: 0–50, including NH3: 0–50 | – | High | Available | Applicability to grain-supplemented systems | Christie et al. (2014) and Smith et al. (2021) | |
| Nitrification-inhibitor urine-patch sprays | – | Direct N2O: 0–80 NO3− leaching: 30–50 | – | High | Available | Cost; regulation of DCD; practicalities of on-farm use | Di and Cameron (2002), Kelly et al. (2008), Di and Cameron (2016) and Ward et al. (2018) | |
| Plantain pastures | – | Direct N2O: 30 NO3− leaching: 20–60 | – | Medium | Available | Integration of plantain into current Australian pasture systems | Box et al. (2017), Minnée et al. (2020), Rodríguez Gelós (2020), Vi et al. (2023), Eady et al. (2024) and Fransen et al. (2024) | |
| On-off grazing | – | Direct and indirect N2O: 7–11 | – | Low | Available | Most applicable in wet climates | de Klein et al. (2006) and Luo et al. (2013) | |
| Covered anaerobic ponds | – | – | Total farm CO2e: 5 | High | Available | Not economically feasible except for very large farms | Craggs et al. (2008) and Grell et al. (2024) | |
| Flocculation | – | – | Waste CH4 from dairy yard effluent: 85–99 | Low | Available | Number of studies; capital costs | Cameron and Di (2021) and Grell et al. (2023) | |
Enteric methane (CH4), direct and indirect nitrous oxide (N2O) and waste CH4 abatement measures are provided where available. Other measures are provided in their absence, including abatement of nitrogen (N) losses, N fertiliser, direct N2O, ammonia (NH3), nitrate (NO3−) leaching or total farm carbon dioxide equivalents (CO2e). Certainty is based on the number of peer-reviewed studies and whether a statistically significant effect on GHG abatement has been established. High certainty has been assigned where more than 15 peer-reviewed in vivo studies have been conducted and have shown a relationship with GHG abatement including in grazing-based dairy systems. Medium certainty has been assigned where five or more peer reviewed studies show a relationship with GHG abatement. Low certainty has been assigned where fewer than five peer-reviewed studies have shown a relationship with GHG abatement.
DCD, dicyandiamide. 3-NOP, 3-nitrooxypropanol.
With technology available now, enteric CH4 on Australian dairy farms could theoretically be reduced by 40–50% with a CH4 inhibitor, feeding lipid by-products and breeding, and waste-CH4 emissions from effluent storage could be abated by another 5–10% of farm CO2e emissions by use of covered anaerobic ponds or flocculation. Reductions in urine-derived emissions could be achieved with nitrification inhibitors to treat urine patches, improving the dietary energy to protein ratio and incorporating plantain into pastures, whereas N fertiliser-derived N2O emissions can be reduced through lowering N fertiliser and treating fertiliser with nitrification inhibitors, but results are highly variable and abatement estimates are subject to local conditions.
However, in practice, there are significant barriers to emissions reductions. Few strategies are currently cost-effective for farmers, limiting an incentive for farmers to adopt strategies. Although industry analysis identifies the relative cost-effectiveness of many abatement strategies, (Dairy Australia 2023c), there is a lack of information on how costs can be distributed for adoption to occur. This highlights the importance of economic research to determine the most cost-effective methods to encourage adoption of abatement methods by dairy farmers. Supply-chain analysis is required to understand how costs could be distributed equitably, and investigation of how a new shared business model could be developed is recommended. Potential government policy support should also be investigated.
Significantly more research investment is required to facilitate the on-farm adoption of abatement strategies. Australian research should be prioritised that can give Australia’s mostly grazing-based dairy systems a competitive advantage. Research must focus on applicability and practicality of abatement methods in grazing systems and build on international research in grazing-based dairy systems. Because enteric CH4 makes up the largest proportion of emissions, reductions in enteric CH4 should be a high priority. Investment in methods that facilitate the use of CH4 inhibitors and rumen modifiers in grazing-based systems will be essential to matching the emissions reductions that will be possible in countries in which confined dairies are prevalent. This should include investment in less well-developed methods, such as early-life programming, that could be particularly relevant and cost-effective where supplements cannot be provided in every mouthful of feed. Methods for abating N emissions should also have a high priority because of the importance of pasture in Australian-based systems. Additionally, most strategies have demonstrated low to moderate abatement, and many require improved certainty around their emission-abatement potential if they are to be adopted. Therefore, a greater understanding is required of the mechanisms behind abatement to improve understanding of abatement potential, possible abatement combinations, interactions with economically beneficial traits and animal- and human-health implications. Such research investment and adoption of cost-effective emission-reduction methods applicable to grazing systems would bolster the position of Australian dairy in the international supply chain.
Acknowledgements
The authors thank Stephen Garnett and Gabriel Crowley for their comments on this manuscript.
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