Effects of water-quality management mitigations on greenhouse-gas emissions from deer farms

Keywords: air, bedding, feed intake, forage type, greenhouse gas, red deer, water, wintering systems.

Environmental degradation remains a significant concern when the utilisation of grasslands intensifies (Merten and Minella 2013). New Zealand grazing systems, evolved over 150 years, have captured significant gains in productivity (Fennessey et al. 2016), although are now showing signs of stress (Chapman et al. 2022). Red deer farming, initiated in the 1970s, has followed this trend as significant markets for venison and velvet antler have been developed (Gray 2021). In recognition of these challenges to the environment, regulatory frameworks to control emissions to water and air are being implemented in New Zealand (McFarlane et al. 2020).

Outdoor wintering systems dominate management practices throughout New Zealand. Forage cropping, using brassicas and fodder beet, provides significant amounts of the feed requirements of livestock, especially in the South Island of New Zealand. Typically, between 3% and 7% of farmland is dedicated to winter forage crops on intensive grazing farms (Stevens et al. 2021). However, wintering on forage crops can have a significant impact on nutrient loss (Smith and Monaghan 2020) and soil damage, resulting in sediment loss to waterways (Monaghan et al. 2017). Farmers are responding quickly, exploring options that minimise this impact.

Red deer have a range of behavioural characteristics that exacerbate potential for sediment loss and bacterial contamination of water ways. Fence-line pacing is known to decrease pasture cover and increase erosion risk (McDowell et al. 2004). Wallowing in water, a natural behaviour, also increases potential phosphorus and bacterial loadings in waterways (McDowell and Stevens 2006). Various managements have been proposed to mitigate these activities (McDowell and Stevens 2006), leaving winter management of forage crops as a significant on-going source of contaminants when deer farming (McDowell and Stevens 2008), similar to other outdoor ruminant grazing systems in New Zealand (McDowell et al. 2003).

Wintering systems are a significant source of contaminants to waterways because stock graze at a time of year when drainage or surface runoff occurs. Both runoff and subsurface flows transport contaminants off land and into nearby waterways (Monaghan et al. 2017). Deer grazing both pasture and forage crops during winter contribute to contaminant losses (McDowell and Stevens 2008). While the loss of sediment from forage crops (1010 kg/ha) was approximately 30 times greater than losses from pasture (31 kg/ha), the total amount lost per farm will depend on the percentage of pasture or forage cropping on any farm. Potential losses ranged from 20 to 4480 kg/ha (McDowell and Stevens 2008), indicating the opportunity for management to control contaminant loss.

Avoiding winter forage crops by bringing livestock indoors is one approach to minimising contaminant losses to waterways. Bringing livestock indoors for winter changes the farm dynamic. Grazing practices are replaced by diet and animal health management, while land use is replaced by capital investment in infrastructure. It also has the potential to alter greenhouse-gas (GHG) emissions as home-grown feeds are substituted for imported feed, adding extra processing and transport, and altering emission factors. For example, the digestion of in situ kale averaged 28% lower methane emissions (Thomson et al. 2016) than did digestion of grass-based diets for sheep. However, nitrous oxide emissions from urine deposited on soil after forage crop grazing by dairy cows are approximately double those from an equivalent pasture grazing event (derived from van der Weerden et al. (2017)).

Farmers must carefully consider the overall impacts of practice change when faced with the need to reduce both contaminant loss to water and emissions loss to air. In this research, we document the potential to prevent contaminant losses to water while comparing different models of GHG emissions for a group of farmers who were aiming to reduce contamination of water by replacing in situ forage crop grazing with winter-housing systems.

Five farmers in the southern region of New Zealand developed indoor facilities (barn wintering) to house herds of deer during winter. These facilities were developed to replace in situ grazed winter forage crops as a source of winter nutrition. This allowed the housing of animals to prevent damage to water-logged and erosion-prone soils from winter forage grazing.

Farmers housed a range of livestock classes, most predominantly velveting stags of a range of ages, but including hinds and weaners. In total, 32 herds were used over a 2-year period. The average herd size was 115 animals, with an average indoor period of 87 days (Table 1). Bedding used was either a bark chip or sawdust deep-litter system, with occasional additions during the winter if deemed necessary. A top-soiled layer (approximately 10–20 cm in depth) was removed at the end of each season and refreshed with new material. Bedding lasts between two and four winters. Animals were housed to meet the animal-welfare guidelines specific for deer (Ministry for Primary Industries 2018).

Table 1.  Farm descriptions, including the type of livestock wintered indoors, herd size and length of winter feeding.

Actual feed intake of the animals housed indoors was calculated from data of feedstock used, including silage, baleage and imported concentrate supplements. Nutritive value of the feed provided (Table 2) was determined by near-infrared spectroscopy (NIRS; Hill Laboratories, Hamilton New Zealand).

Table 2.  Mean dry-matter content, metabolisable energy concentration and N concentration in feeds supplied to red deer fed indoors or outdoors on a kale forage crop during winter.

Feed requirements if each herd was grazed on a forage crop were calculated from equations published by Nicol and Brookes (2007) for New Zealand red deer. These feed requirements were then used, in conjunction with herd numbers and nutritive value of forage crops, to predict the land area required for each herd. A yield of 14 t DM/ha was assumed, on the basis of yields in the region of study published in the literature (Thompson et al. 2010; Smith et al. 2012; Monaghan et al. 2017).

Values of 11.4 MJ ME/kg DM and 152 g CP/kg DM (Westwood et al. 2014) were used in calculations to determine dry-matter intake and potential N output. Nitrogen outputs via urine and faeces were determined by the following process:

1. Nitrogen intake was calculated (dry-matter intake (DMI; kg/day) × N concentration in the feed (g/kg DM))

2. Faecal output was calculated (intake × (1 − (ME/GE/0.81))) using eqns 1.5 and 1.10 (SCARM 1994)

3. Nitrogen output in faeces was calculated (−4.623 + (197 × feed N (g N/g DMI)) + (feed intake × 7.89)) × 0.001 (eqn 5.7 from Ministry for Primary Industries (2022))

4. Nitrogen output in urine was calculated by difference (N intake – N in faeces)

Estimates of losses of sediment, phosphorus (P) and nitrogen (N) were made to indicate the potential reduction in losses by using barn wintering. Monaghan et al. (2017) compared two forage crop grazing management approaches over 2 years in a paired-catchment study on soil types that were similar to those farmed by the study group. They concluded that the average loss of sediment, P and N (in both overland and subsurface flow) was 4130, 4.4 and 39 kg/ha.annum respectively, under a standard grazing regime with dairy cows. Using a strategic grazing approach, which avoided critical source-area grazing until the end of the grazing period, average loss of sediment, P and N (in both overland and subsurface flow) was 830, 1.4 and 21 kg/ha.annum respectively. Given that farmers are progressing towards implementing the strategic grazing approach, the averages of these two sets of values (2980, 2.9 and 30 kg/ha.annum for sediment, P and N respectively) have been used to estimate losses. When grazing forage crops in situ, some wastage is measured. A feed-utilisation value of 85% (Judson and Edwards 2008) was applied to crop yield to determine total area required for grazing, and hence contributing to potential soil loss.

Loss of nitrogen to air as ammonia (NH₄) also occurs from dung and urine deposition. The New Zealand IPCC inventory methodology (Ministry for Primary Industries 2022) has estimated NH₃ emissions to then calculate indirect N₂O emissions. The NZIPCC methodology value of a 10% loss of N to NH₃ is used to estimate losses to air from grazing winter forage crops. Nitrogen deposited as dung and urine to bedding when wintering indoors also creates ammonia. Measurements reported when dairy effluent was mixed with sawdust (van der Weerden et al. 2014) showed that a total of 1.14% and 3.0% of N deposited was lost as N₂O and NH₃ respectively, after 7 months. These values were used to estimate losses to air in this analysis.

Calculations of GHG emissions included carbon dioxide (CO₂), methane (CH₄) and nitrous oxide (N₂O), from production of feed and its transport and from livestock emissions. Emissions from the production and transport of supplementary feed used emission factors published by Ledgard and Falconer (2015) for New Zealand cropping practices (Table 3). These include emissions relating to both fuel use and land-based emissions from cultivation and fertiliser use. Transport emissions were added to reflect delivery of supplementary feed sourced off-farm, at the rate of 100 g CO₂/t.km, assuming an average 100 km round trip from supplier to the farm.

Table 3.  Carbon footprint of supplementary feed and forage crops in New Zealand (from Ledgard and Falconer 2015).

New Zealand runs tier-two country-specific emissions factors for livestock. Methodology from the New Zealand national GHG inventory calculations was used to calculate GHG emissions (Ministry for Primary Industries 2022). Briefly, this allocates GHG emissions on the basis of enteric methane production from digestion, methane from dung degradation, and N₂O from dung and urine deposition to land, using specific factors of N loading in dung and urine (van der Weerden et al. 2020). Standard values and equations specific for deer (Ministry for Primary Industries 2022) were used to calculate emissions from digestion (methane), dung deposition (methane and N₂O) and urine deposition (N₂O) (Table 4). Adjustments were made to these calculations to reflect variations due to forage crop digestion and soil conditions under winter forage crop grazing.

Table 4.  Values for methane production and nitrous oxide loss used to calculate greenhouse-gas emissions.

A meta-analysis of measurement of methane emissions from the enteric digestion of forage crops (Thomson et al. 2016) reported changes in methane emissions from +13% to −64%. The average change in methane emissions was −28% from both sheep and cattle. The methane emissions (g CH₄/kg DM consumed) determined by the standard calculations were reduced by this amount (Table 4) for deer fed forage crops.

Nitrous oxide emissions from soil increases when saturated and disturbed by forage crop grazing (van der Weerden and Styles 2012; Monaghan et al. 2013; Di et al. 2016; Treweek et al. 2016; van der Weerden et al. 2017). To reflect the greater emissions from disturbed exposed soils post-forage grazing, the loss of N deposited as urine was adjusted as per results documented in Table 4. The average N to N₂O loss of N deposited to forage crop soil was of 1.94% from dairy cows. This is two times that emitted from urine of cattle deposited on pastures. Using this logic, the emissions from deer urine deposited to soils during forage crop grazing was doubled from 0.74% N lost as N₂O for pasture to 1.48% N lost as N₂O for forage crops (Table 5).

Table 5.  Literature values for nitrous oxide emission factors from dairy cattle grazing winter forage crops.

These were converted to CO₂ equivalents using current Global Warming Potential (GWP) factors, as per animal emissions (Ministry for Primary Industries 2022).

Data were summarised to compare standard inventory calculations with calculations customised for the wintering barn or forage crop situations. Contaminant results are presented at the herd level for wintering method (barn or forage crop) and stock type (rising 2-year-old stags, rising 4-year-old stags, mixed-age stags, mixed-age hinds, mixed-sex weaners). GHG emissions are presented at the individual animal level.

Estimates of individual intake of red deer in barns, from records of feed stockpile use, increased with animal size (Table 6). Calculated potential forage crop intakes were often numerically greater than the barn recorded intake, averaging 0.26 kg DM/day or 9.5% more than barn intakes. This difference was greatest for mixed-age stags at 0.95 kg DM/day. Nitrogen intakes varied at the same level of magnitude as dry-matter intake, although were 20% greater on the forage crop, due to higher dry-matter intake (Table 6) and higher N content (Table 2).

Table 6.  Estimates of feed and nitrogen intake in barn-fed and forage-fed red deer of a range of genders and ages.

Estimates of potential contaminant loss when herds graze a forage crop were proportional to the area required for grazing (Table 7). Therefore, mixed-age stags, with both the greatest calculated intake and the largest average herd size, incurred the greatest estimated losses. Nitrate losses and ammonia losses were approximately equal and were both an order of magnitude greater than was P loss (Table 7).

Table 7.  Estimates of grazing area required and losses of P, nitrate, ammonia and sediment from soils under forage crops grazed by red deer during winter.

The calculation of methane emissions from enteric fermentation and from dung (Table 8) demonstrated the differences that may occur as emission factors, intake and digestibility vary. For example, methane emissions from dung were influenced by digestibility of the diet, as faecal output was reduced on the forage crop, resulting in lower emissions from dung (Table 8). Methane from enteric digestion was calculated to be approximately 10% greater when red deer were wintered on a forage crop when inventory values were used, due to a higher intake. However, when the revised value was used, deer fed forage crops had calculated methane emissions that averaged 21% less than those for the deer fed indoors.

Table 8.  Estimates of methane emissions from barn-fed or forage-fed red deer using standard inventory calculations or revised calculations.

Calculated nitrous oxide emissions from dung were approximately 10% greater from the forage crop than were the barn inventory calculations (Table 9). However, when revised to reflect greater potential loss from the interaction with the sawdust bedding, the calculated direct N₂O emissions from dung were an order of magnitude greater for the barn system (Table 9). The revised barn calculation was 65% more than the inventory calculation for urine N₂O losses (Table 9). The revised forage crop calculation was 100% greater than the inventory calculation for urine N₂O losses (Table 9).

Table 9.  Estimates of nitrous oxide emissions from barn-fed or forage-fed red deer using standard inventory calculations or revised calculations.

Indirect N₂O losses from NH₃ volatilisation were numerically similar when calculated using the inventory calculation. The revised barn calculation resulted in a reduction in indirect N₂O losses from NH₃ volatilisation by approximately 68% (Table 9).

Total N₂O losses using the inventory calculations were 22% greater on the forage crop than in the barn (Table 9). However, the revised calculations estimated that the forage crop emissions were approximately 5% lower than were barn emissions. Both revised barn and forage crop emissions were approximately 100% greater than inventory estimates.

Greenhouse-gas estimates from animal activities by inventory calculation were higher from the forage crop than barn system (Table 10). When revised to reflect changes due to differences in enteric CH₄ emissions from forage crops and N₂O and NH₃ losses, emissions from deer housed in barns were 11% higher than inventory calculations, while those from deer fed on forage crops were approximately 16% lower than inventory calculations.

Table 10.  Estimates of total greenhouse-gas emissions from barn-fed or forage-fed red deer using standard inventory calculations or revised calculations.

Greenhouse-gas emissions from the provision of food to the barn or forage crop deer were similar (Table 10). Total GHG emissions followed the same pattern as those produced by animal activities, with the revised calculations being higher than inventory calculations for the barn system (+8%), while being lower for the forage crop system (−12%).

Calculated contaminant losses to the environment were considerable when forage crops were fed to deer during the winter. Larger animals such as mixed-age stags with greater daily feed requirements had a larger impact through requiring a greater area for food provision. Various researchers have documented the extent of losses when red deer graze (e.g. de Klein et al. 2003; McDowell and Stevens 2006, 2008), ranging from 440 to more than 2200 kg soil/ha.annum, with losses from winter crop grazing estimated at approximately 1000 kg soil/ha.annum (McDowell and Stevens 2008). The estimated values in the current study were greater than those reported in the literature. This may be related to several factors that influence potential contaminant loss. These include the type of animal used, the duration of grazing/soil disturbance and the soil type itself.

The use of dairy cattle data to provide the estimates may be influenced by the size of the animal, and the damage done by the downward pressure applied to the soil. Soil disturbance by the hooves of the grazing animal is a major contributor to potential soil loss and is much greater with cattle than sheep (Donovan and Monaghan 2021). For example, the downward pressure applied by a 670 kg dairy cow with a hoof area of approximately 50 cm² has been measured to be 180–200 N/cm² at push-off during walking (van der Tol et al. 2003). During this action, this relates to the transfer of approximately half of the weight of the animal to the area of hoof still in contact with the ground. van der Tol et al. (2003) measured this area to be approximately one-third of the resting area of the hoof, or 15 cm². When translated into metrics for red deer, a stag has an estimated hoof area of approximately 36 cm² and a liveweight of approximately 300 kg. Using the relative data from van der Tol et al. (2003), this would translate into a force of approximately 120–130 N/cm², just over half that exerted by the dairy cows reported by van der Tol et al. (2003). The liveweight of the red deer in this study typically sits between dairy cows and sheep, leading to the potential to result in lower soil losses when grazing forage crops than the dairy cow estimates that were used here.

Duration of grazing, or the length of time the animals walk over the soil, also affects potential contaminant loss, as the number of hoofprints increases. The data reported by McDowell and Stevens (2008) from deer represented a 2-week grazing period, while Monaghan et al. (2017) reported on an 8-week period of dairy cattle grazing, potentially increasing soil damage by four-fold.

Soil type, too, can have a significant impact on soil loss under intensive winter grazing due to relative susceptibility to physical degradation (Hewitt and Shepherd 1997). The Pallic soil grazed by dairy cattle (Monaghan et al. 2017) are known to be structurally vulnerable (Hewitt and Shepherd 1997) and have been recorded to have a resistance strength of approximately 0.2–0.3 MPa when saturated during winter crop grazing (Thompson et al. 2010), which is approximately 10% of the downward force of the dairy cow reported by van der Tol et al. (2003). This confirms the vulnerability of these soils to damage during the grazing of a winter forage crop, increasing the potential for contaminant loss (Donovan and Monaghan 2021). The soils in the deer study reported by McDowell and Stevens (2008) were, by the description of the site, most likely Brown soils, which are much more resistant to physical damage (Hewitt and Shepherd 1997).

Contaminant loss can be reduced using good grazing management practices to ensure that soil damage is minimised through lower livestock density (offering multiple days of feed allocation at a time), and/or through minimising time spent on each allocation of land (daily shifting to a new forage area for example). These reduce the relative grazing intensity and result in reduced soil damage (Donovan and Monaghan 2021). Measures to prevent grazing of critical source areas also reduce potential soil losses from winter forage crop grazing (Monaghan et al. 2017).

Overall, calculated GHG emissions are similar to those in other literature. Most reports are from work conducted for whole years, rather than short periods such as this. However, comparison can be made using the yield of GHG per kg DM eaten. For example, the output of animal GHG emissions from dry-matter intake using the standard calculation method in this study is approximately 0.65 kg CO₂-e/kg DM, similar to the 0.67 kg CO₂-e/kg DM that can be calculated from data reported by Vibart et al. (2021) for low-intensity sheep and beef farming.

More intensive dairy farming systems include more supplementary feed and winter forage cropping, increasing the contribution from sources such as cultivation and the mechanical processes associated with this and supplementary feed conservation. The total GHG output calculated for dairy cattle from van der Weerden et al. (2018) is 0.8–0.85 kg CO₂-e/kg DM. Inventory calculations provided values of 0.87 and 0.89 kg CO₂-e/kg DM for barn and forage crop systems respectively. The revised calculations of the current data set were 0.96 and 0.77 kg CO₂-e/kg DM, for barn and forage crop systems respectively. While these values are higher than those calculated from data reported by van der Weerden et al. (2018), the values in the current data set represent only 87 days, or 24% of the year. These 87 days are the most intensive, so the expectation is that transitioning back to a pasture-based feeding regime would result in a decrease in intensity. Assuming the remainder of the year had outputs similar to Vibart et al. (2021), an overall GHG output of approximately 0.70 kg CO₂-e/kg DM would be the result.

Both revised calculations of N₂O emissions were nearly double that predicted by the inventory calculations. Soil conditions have a significant role in altering N₂O emissions, especially from urine deposition. Compaction increased N₂O emissions from 0.3 to 0.75 kg N₂O/kg N applied (van der Weerden et al. 2017). This may be soil related as the opposite has occurred when soil compaction is very high, although not with urine deposition, and the emissions factor was 3.2 kg N₂O-N/kg N (van der Weerden and Styles 2012). Smith et al. (2008) measured an emissions factor of 1.4 N₂O-N/kg N from urine deposited during the grazing of swedes. All may depend on extent of compaction and anaerobic conditions. Prolonged treading damage on saturated soils can cause failure of the soil structure leading to anaerobic conditions in some forage grazing situations (McDowell et al. 2003). Emissions from cattle dung deposition average 0.005 kg N₂O/kg N applied (van der Weerden et al. 2021). Incorporation into the soil will occur with crop use but did not make any difference to N₂O emissions, leading to the standard application of an emission factor of 0.006 by the IPCC.

Ammonia from cattle manure applied as a slurry is also included in international inventories (van der Weerden et al. 2021). Incorporation into the ground (crop) has an emissions factor of 0.129, while the emissions factor for broadcast application is 0.242. If it remains a solid, then the NH₃ emission factor is reduced to 0.03. These large variations in potential emissions may have significance to values calculated on-farm when moving from outdoor to indoor systems. The New Zealand inventory calculations may be lacking some elements to adequately assess the GHG from barn systems as barn manure emission factors are not presently included. Emissions from sawdust-based bedding altered the pathway of emissions (van der Weerden et al. 2017).

Methane emissions were reduced when forage type was accounted for. Thomson et al. (2016) summarised nine experiments that compared a range of winter forage types, including kale, rape, turnip, swede and fodder beet, with control diets, mainly of ryegrass pasture. A reduction of methane emissions was recorded in 27 of 28 treatments feeding winter forages. These ranged from −2.4% to −61% compared with the control and were related most closely to the dietary NDF concentration (R² = 0.40). This provides some confidence that the emission factor applied to winter forages should be lower than that for the mainly pasture-based supplements used in the barn feeding systems.

The data presented here demonstrated some of the variability that may occur both in contaminant loss to water and emissions to air. It also provided an indicator of the potential difference in GHG emissions that may be possible as knowledge of the emission processes are refined.

The data that support this study will be shared upon reasonable request to the corresponding author. 

The authors declare no conflicts of interest.

Deer Industry New Zealand and AgResearch provided funding for this research.