Composting organic matter can lower the global warming potential of food and agricultural waste and provide a nutrient-rich soil amendment. Compost applications generally increase net primary production (NPP) and soil water holding capacity and may stimulate soil carbon (C) sequestration. Questions remain regarding the effects of compost nitrogen (N) concentrations and application rates on soil C and greenhouse gas dynamics. In this study, we explored the effects of compost with different initial N quality (food waste versus green waste compost) on soil greenhouse gas fluxes, aboveground biomass, and soil C and N pools in a fire-impacted annual grassland ecosystem. Composts were applied annually once, twice, or three times prior to the onset of the winter rainy season. A low intensity fire event after the first growing season also allowed us to explore how compost-amended grasslands respond to burning events, which are expected to increase with climate change. After four growing seasons, all compost treatments significantly increased soil C pools from 9.5 ± 0.9 to 30.2 ± 0.7 Mg C ha^-1 (0-40 cm) and 19.5 ± 0.9 to 40.1 ± 0.7 Mg C ha^-1 (0-40 cm) relative to burned and unburned controls, respectively. Gains exceeded the compost-C applied, representing newly fixed C. The higher N food waste compost treatments yielded more cumulative soil C (5.2 to 10.9 Mg C ha^-1) and aboveground biomass (0.19 to 0.66 Mg C ha^-1) than the lower N green waste compost treatments, suggesting greater N inputs further increased soil stocks. The three-time green waste application increased soil C and N stocks relative to a single application of either compost. There was minimal impact on net ecosystem greenhouse gas emissions. Aboveground biomass accumulation was higher in all compost treatments relative to controls, likely due to increased water-holding capacity and N availability. Results show that higher N compost resulted in larger C gains with little offset from greenhouse gas emissions and that compost amendments may help mediate effects of low-intensity fire by increasing fertility and water holding capacity.

Methods

Field site

The site was located at the University of California Sierra Foothill Research and Extension Center (SFREC) in Browns Valley, California. Soils are derived from Mesozoic and Franciscan volcanic rock and classified as xeric Inceptisols and Alfisols in the Auburn-Sobrante complex (Soil Survey Staff 2020). The site had been grazed by cattle for at least 150 years (D. Flavell, personal communication). Average annual precipitation was 700 mm with pastures producing on average 3,300 kg ha^-1 y^-1 of biomass dominantly used for livestock forage. The study region has a Mediterranean climate where the growing season typically occurs from October to April or May and is characterized by cool, wet winters and warm dry summers.  The site was dominated by naturalized stands of annual grasses and forbs (Bartolome et al. 2007, Eviner 2016). The field sites were not seeded, irrigated, fertilized, or tilled. Naturalized annual plant species reseed and replace stands every year. This ecosystem is broadly representative of approximately 12 million hectares of rangeland across California (Eviner 2016).

Table 1 Description of soil amendment, application, and burned or unburned control treatments.

Experimental design

In October 2018, nine original 0.15 ha (60 x 25 m) plots were established with 6 m buffers to establish a randomized complete block design with three complete blocks. Each block contained a food waste compost treatment with applications in two years (FW or FW2), a manure and green waste compost treatment with applications in three years (hereafter referred to as green waste compost and GW, GW2, or GW3), and an untreated burned control (UCN); an unburned control (CN) was added in spring 2019 (Table 1). The food waste compost had higher mineral N (80 µg g N^-1) than the green waste compost (57 µg g N^-1), but the composts had similar total C and N (Tables S1-S2). To apply the second compost treatments in Fall 2020 (FW2, GW2), original plots, except the controls, were split into 0.075 ha (60 x 12.5 m) plots for all treatments (n = 3 per treatment for CN, FW, FW2, GW, GW2). To apply the third compost treatments in Fall 2022 (GW3), the GW2 plots were split in half again to create 0.0375 ha (30 x 12.5 m) plots for GW2 (n = 3) and GW3 (n = 3) treatments.  In June 2019 shortly after soil and plant samples were collected, a fast low intensity grass fire burned all plots evenly. Following the burning event, unburned controls (UCN) 0.15 ha (60 x 25 m) plots were established approximately 80 m from the closest burned plot and sampled identically thereafter to further elucidate the effects of fire across treatments. These treatment groups allowed us to simultaneously compare the effects of green waste versus food waste compost, the effects of compost application frequency, and the effects of fire on a composted grassland ecosystem. Fire is a natural and common event in these grasslands (Harrison et al. 2003).

Established plots were located on similar slope and aspect; management and soil type (pre-treatment C:N of 12)  were consistent across the whole study area. Treatments were randomly assigned within each block. Both green and food waste composts were produced at the West Marin Compost Facility (for compost properties see Tables S1-S2) by maturing in watered piles that were turned (aerated) weekly for three months (Vergara and Silver 2019, Pérez et al. 2023). Food waste compost (C:N of 16) was derived from food scraps including fruits, vegetables, egg and clam shells, bones, etc. Green waste compost (C:N of 16.4) was derived from plant residues, chicken, horse, and cattle manures. Both feedstocks were mixed with woodchips in a matrix that was 50% by volume to facilitate proper compost development and maturation. All plots were grazed by yearling steers for approximately three weeks and then mowed to a uniform aboveground cover prior to initial compost application. During grazing, cattle were allowed to graze all plots freely and not isolated to any specific plot. Composts were applied at a depth of 0.65 cm (equivalent to 5.9 Mg C ha^-1 / 0.37 Mg N ha^-1and 5.5 Mg C ha^-1 / 0.34 Mg N ha^-1 for food waste and green waste compost per application, respectively) to respective treatment plots in November 2018, October 2019, and October 2021 using a compost spreader (application dates based on best practices by the range manager). Control plots were driven over without amendment application to impart the same soil effect.

Soil sampling and analyses

Soil sampling was conducted prior to compost application (fall 2018) and annually at the end of each growing season (end of spring) in 2019, 2020, 2021, and 2022. At every sampling time point, soil samples (n = 9 per plot regardless of plot size) were collected at 0-10 and 10-20, and 20-30 cm depths from nine stratified random locations per plot using a soil auger. When possible, soil samples were also collected from 30-40 and 40-50 cm depths. Due to the inability to sample 40-50 cm in some locations, these depths were not included in soil C stock calculations but are provided in the Supplementary Information (Table S3). All compost addition subplots were sampled at least 5 m from the edge of each subplot boundary to minimize edge effects between treatments with application frequency. Samples were transported to the laboratory and processed for analysis within 24 hours. Soil pH was measured in a 1:1 volumetric slurry of sample and deionized water using a pH electrode (Mclean 1982). Soil moisture was determined gravimetrically by weighing fresh soil and oven-drying for at least 24 hr at 105 ^ºC. Mineral N species and N mineralization rates were quantified throughout the first growing season (early, mid late growing season, and end of growing season) and during the last annual soil sampling (end of growing season 2022). Nitrate plus nitrite (NO₂^-) and NH₄⁺ were measured after extraction of 15 g of field-fresh soil in 75 ml of 2 M potassium chloride (KCl) solution (Hart et al. 1994). Potential net nitrification and mineralization were measured by comparing fresh 2M KCl soil extractions with a second subsample was covered and incubated for seven days in the dark, prior to subsequent 2M KCl soil extractions. Soil KCl extracts were stored at -20 ^ºC until colorimetrically analyzed using an AQ300 analyzer (Seal Instruments, Mequon, WI). The difference in NO₃^- over time was used to calculate potential net nitrification, and the difference in the sum of NO₃^- and NH₄⁺ concentrations over time was used to calculate potential net mineralization, after accounting for the length of the incubation (Hart et al. 1994).

For total soil C and N analyses, subsamples were air-dried, sieved to < 2 mm, and had visible roots removed before being ground to a fine powder. Samples were then analyzed in duplicate for total C and N on a CE Elantech elemental analyzer (Lakewood, NJ). Bulk density was sampled in 2020 from three locations in each plot using a 6.5 cm diameter bulk density corer. Samples were sorted in the lab into fine soil (< 2 mm) and coarse rock (> 2 mm) volumes. To quantify soil moisture, soil subsamples from each depth and replicate were weighed before and after drying at 105 ^ºC to a constant weight for at least 24 h. Bulk density was calculated as the rock-free dry volume for total soil fractions, and used to calculate total C and N stocks (g C m^-2, g N m^-2). Previous work at this site suggested that bulk density was relatively similar following compost amendments over one-to-four-year time scales (Ryals et al. 2014).

Soil greenhouse gas fluxes

Fluxes of CO₂, N₂O, and CH₄ were measured across the soil-atmosphere interface using the static chamber method in years one and three; they were not measured in year two (2020) due to complications from the COVID-19 pandemic. Three chambers were sampled in each plot. Gas sampling took place at daily intervals after the initial compost application during the spring wet up period for 8-21 days, and over the entire growing season at bi-weekly intervals. Gas samples (30 mL) were collected from each chamber at 0, 5, 15, 25, and 40 minutes after the lid closure and stored in pre-evacuated gas vials until analysis (within 72 hours) for CO₂, N₂O, and CH₄ on a Schimadzu GC-14 gas chromatograph (Shimadzu Corporation, Kyoto, Japan) equipped with a thermal conductivity detector, flame ionization detector and an electron capture detector. Analyzer detection limits were 0.09 ppm CH₄, 0.49 ppb for N₂O, and 0.09 ppm for CO₂. Fluxes were calculated using an iterative exponential curve-fitting approach with non-statistically significant fits considered as fluxes equal to zero (Matthias et al. 1978, Ryals and Silver 2013).

Aboveground biomass

Aboveground vegetation was sampled for biomass production and C and N content at peak standing crop at the end of each growing season (n = 9 samples per plot per growing season). Aboveground shoots were clipped from 20 cm diameter circles, collected into pre-dried bags, dried at 65 ºC for > 24 h, and subsequently weighed. Dried and ground subsamples were analyzed for C and N on a CE Elantech elemental analyzer (Lakewood, NJ).

Statistical analyses

Response variables were analyzed using JMP 16 (SAS Institute Inc., Cary, NC). Soil and biomass data were analyzed using a linear mixed model with block as a random effect and treatment (green and food waste composts and control) and time (2019: year one, 2020: year two 2021: year three, and 2022: year 4) as fixed effects. For significant treatment effects on soil C and N, plant productivity, and greenhouse gas fluxes, all treatment groups (FW, FW2, GW, GW2, GW3) and control groups (CN, UCN) means were compared using a Tukey’s post-hoc test. As aboveground plant biomass exhibited high interannual variability, biomass data were further assessed for significant trends using mean change in plant biomass (treatment-control) and applying a one-sample t-test for the green waste compost treatment and for food waste compost. Values reported in the text are means and standard errors.