Catena 198 (2021) 105046 Available online 19 November 2020 0341-8162/Crown Copyright © 2020 Published by Elsevier B.V. This is an open access article under the CC BY-NC-ND license (http://creativecommons.org/licenses/by-nc-nd/4.0/). Nutrient retention, availability and greenhouse gas emissions from biochar-fertilized Chernozems Carlos M. Romero a,b,*, Xiying Hao b, Chunli Li a,b, Jen Owens b, Timothy Schwinghamer b, Tim A. McAllister b, Erasmus Okine a a Department of Biological Sciences, Faculty of Arts & Science, University of Lethbridge, 4401 University Drive, Lethbridge, Alberta T1K 3M4, Canada b Agriculture and Agri-Food Canada, Lethbridge Research and Development Centre, 5403 1st Ave. S., Lethbridge, Alberta T1J 4B1, Canada A R T I C L E I N F O Keywords: Biochar Greenhouse gases Soil nutrients Chernozems Canadian Prairies A B S T R A C T Amending soil with pyrogenic-C (biochar) has emerged as a potential best management practice to accumulate organic matter (OM), reduce greenhouse gas (GHG) emissions and increase nutrient retention among degraded, marginally-productive croplands. Nevertheless, the impact of biochar application on intensively cropped prairie eco-regions is not well documented, particularly when co-applied with chemical fertilizer. Our objective was to determine the effect of biochar on cumulative CO2-C, N2O-N and CH4-C emissions, water-extractable OM, and available N (AN; NH4-N + NO3-N) and P (AP; PO4-P) in the presence or absence of NP-fertilizer. Biochar was applied to two surface Chernozems (0–15 cm) of contrasting texture [i.e., sandy clay loam (Raymond) and clayey (Lethbridge)] at six rates (0, 0.15, 0.5, 3, 10 and 20 Mg ha− 1) in combination with (+NP) or without (− NP) urea- N (150 kg N ha− 1) and KH2PO4 (50 kg P ha− 1). A total of 72 soil columns were incubated at 21 ◦C for 86 d. Biochar addition increased soil total C up to 24.9 g kg− 1 and 28.7 g kg− 1 in Raymond and Lethbridge, respec- tively, but did not affect water-extractable OM. Cumulative N2O-N and CH4-C emissions were not influenced by biochar, regardless of whether or not NP-fertilizer was added (p > 0.05). Cumulative CO2-C emissions varied between soil textures and were increased or decreased non-linearly by biochar addition under –NP only. Available P increased within +NP soil with increasing biochar rates reaching 43.9 mg kg− 1 in Raymond and 79.5 mg kg− 1 in Lethbridge when biochar was applied at 20 Mg ha− 1. A similar but less pronounced response was observed for AN. Our results indicate that biochar-only application is not a practical management approach for improving soil fertility and nutrient cycling in surface Chernozems. Nevertheless, co-applying biochar with NP- fertilizer appears to improve soil P availability in the short-term. 1. Introduction Biochar is a form of black carbon (C) produced by O2-limited thermal treatment of animal- or plant-derived organic matter (OM) at tempera- tures <700 ◦C (Hagemann et al., 2017). Amending soil with biochar has been shown to improve soil fertility and environmental quality in agri- cultural (Ippolito et al., 2016; Nelson et al., 2011; Xiu et al., 2019), forest (Kanthle et al., 2018; Pokharel et al., 2018) and grassland ecosystems (Han et al., 2016; Pokharel et al., 2018). Biochar is biochemically recalcitrant (Zimmerman, 2010) and its presence in soil can mitigate agricultural CO2 emissions (Case et al., 2012) by sequestering humified, stable pools of OM (Lehmann et al., 2006) with mean residence times >200 yr (Kuzyakov et al., 2009). Incorporation of biochar into soil can increase soil aeration, and as a consequence, reduce N2O and CH4 production in croplands (Kammann et al., 2017). Amending soil with biochar may also lead to reductions in N2O and CH4 emissions by contributing to increased nutrient retention (El-Naggar et al., 2019a). For example, biochar can adsorb NH4 + (N2O) or CH4 into its porous surface (Jeffery et al., 2016; Cayuela et al., 2014; Nguyen et al., 2017) and immobilize NO3 − (N2O) through higher C/N ratios (Zheng et al., 2012). The influence of biochar on soil nutrient retention has long been established in tropical Anthrosols such as Terra-Petra (Lehmann et al., 2003). Both laboratory and field studies have provided evidence of the liming potential and agronomic benefits of biochar among highly weathered, acidic profiles (<5 pH) (Glaser et al., 2002). For example, * Corresponding author at: Department of Biological Sciences, Faculty of Arts & Science, University of Lethbridge, 4401 University Drive, Lethbridge, Alberta T1K 3M4, Canada. E-mail address: carlos.romero@uleth.ca (C.M. Romero). Contents lists available at ScienceDirect Catena journal homepage: www.elsevier.com/locate/catena https://doi.org/10.1016/j.catena.2020.105046 Received 5 March 2020; Received in revised form 19 October 2020; Accepted 9 November 2020 mailto:carlos.romero@uleth.ca www.sciencedirect.com/science/journal/03418162 https://www.elsevier.com/locate/catena https://doi.org/10.1016/j.catena.2020.105046 https://doi.org/10.1016/j.catena.2020.105046 https://doi.org/10.1016/j.catena.2020.105046 http://crossmark.crossref.org/dialog/?doi=10.1016/j.catena.2020.105046&domain=pdf http://creativecommons.org/licenses/by-nc-nd/4.0/ Catena 198 (2021) 105046 2 Major et al. (2010) reported that soil pH, extractable Ca+2 and extractable Mg+2 concentrations increased after applying 20 Mg ha− 1 of (wood-derived) biochar to a Colombian Oxisol. Biochar amendments also increased corn (Zea mays L.) yields by 28–140% due to reductions in Al+3 toxicity (Major et al., 2010). The application of rice (Oryza sativa L.) straw or leucaena [L. leucocephala (Lam.) de Wit] wood biochar reduced cumulative N2O emissions and restored nitrification activity in sub- tropical Oxisols (He et al., 2016), and increased the physical structure (e. g., aggregate stability), microbial activity (i.e., microbial biomass car- bon) and cation exchange capacity of highly weathered Ultisols (Jien and Wang, 2013). The application of biochar to fertile soils of prairie eco-regions is a relatively new management alternative that has yielded mixed results. For example, Wu et al. (2013) applied 10 and 25 Mg ha− 1 of wheat (Triticum aestivum L.) straw biochar to an incubated Orthic Black Cher- nozem from southern Alberta and noted reductions in cumulative N2O emissions. A similar response was observed by Pokharel et al. (2018), who amended an Orthic Black Chernozem with Korean pine (Pinus koraiensis Siebold & Zucc) sawdust biochar at 17 Mg ha− 1. However, Cheng et al. (2012) reported that the addition of (wheat straw) biochar to a Black Chernozem from north-central Alberta did not influence short- term N2O emissions. In Saskatchewan, Ahmed and Schoenau (2015) applied wheat or flax (Linum usitatissimum L.) straw biochar to Brown or Black Chernozems but found no differences in available N and P, and minor effects on soil pH, electrical conductivity (EC) and OM contents. Co-applying biochar with external nutrient inputs such as NP- fertilizer is now considered a viable approach to mitigating potential biochar-induced reductions in crop yields (Kloss et al., 2014). For example, Mete et al. (2015) reported a synergistic effect of sawdust biochar and NPK-fertilizer on soybean [Glycine max (L.) Merr.] yields within an alkaline soil from Bangladesh. Combined increases in total biomass and seed production resulted from higher nutrient availability, particularly Olsen P (Mete et al., 2015). Nevertheless, Kloss et al. (2014) reported that woodchip biochar could inhibit mustard (Sinapsis alba L.) and barley (Hordeum vulgare L.) growth (by 68%), despite external N- fertilizer inputs (40 or 100 kg N ha− 1) within an alkaline Austrian Chernozem. Although some agronomists have suggested that a single application of biochar (up to 20 or 50 Mg ha− 1) could be a standardized management practice for agricultural purposes, its use has mainly been confined to high-value specialty crops (El-Naggar et al., 2019a). The market price of biochar is high, e.g., ~$1500–2000 USD ton-1 (Jirka and Tomlinson, 2015), making it essential that clear benefits be identified if it is to be routinely used within intensively managed agroecosystems. This is particularly relevant for farm operations in western Canada, where there is a demand for more efficient recycling of non-renewable plant nutrients (Grant and Flaten, 2019). This study aimed to evaluate the effects of six biochar rates, with and without NP-fertilizer, on selected soil properties, cumulative N2O, CO2 and CH4 emissions, and water-extractable OM, N (NH4-N + NO3-N) and P (PO4-P) availability within two surface (0–15 cm) soils from southern Alberta, Canada. We hypothesized that biochar amendment to Cherno- zems would (i) improve soil fertility attributes and (ii) reduce green- house gas (GHG) emissions, in particular when co-applied with external nutrient inputs. 2. Materials and methods 2.1. Characterization of soil and biochar amendment Soil samples were collected from two long-term cropping field sites in southern Alberta, Canada. The study sites were near Raymond (Warner County, 49◦31′12′′N, 112◦42′27′′W; elevation 932 m) and Lethbridge (Lethbridge County, 49◦38′00′′N, 112◦48′00′′W; elevation 929 m). The Raymond soil was classified as a sandy clay loam Orthic Dark Brown Chernozem (Kessler series; Soil Classification Working Group, 1998). The dominant soil at Lethbridge was a clayey Orthic Dark Brown Chernozem (Lethbridge series; Soil Classification Working Group, 1998). Both soils were deep and well-drained and derived from calcareous glaciofluvial (Raymond) or glaciolacustrine (Lethbridge) deposits (CanSIS, 2013). The particle size distribution, based on Gee and Bauder (1986), was: 60.1% sand, 7.6% silt and 32.3% clay for the Raymond soil, and 28.1% sand, 25.4% silt and 46.5% clay for the Lethbridge soil. The Raymond site had been cropped with an irrigated alfalfa (Med- icago sativa L.) – grass mixture since 2014, occasionally grazed by cattle after hay harvest. The Lethbridge site was managed under no-till dryland conditions since 1973 (Hao, 2015) and cropped to a barley- soybean rotation since 2016. Composite samples were collected from the top layer (0–15 cm) using a hand shovel, sealed in tubs, and trans- ported to the laboratory where they were air-dried (1 wk, 22 ◦C), coarsely ground, and passed through a 2 mm sieve. Subsamples of soil (<2 mm) were then finely ground to pass through a 0.15 mm mesh. Initial soil pH and EC were determined using a 2:1 (water:soil) slurry. Finely ground samples were used to determine total C (TC) and total nitrogen (TN) by dry combustion using a CN analyzer (NC2100, Carlo Erba Instruments, Milan, Italy). The total phosphorus (TP) concentration was determined by digesting finely ground samples with 18 M H2SO4 (Parkinson and Allen, 1975). Solutions (digestate) were quantified by colorimetry with a discrete analyzer (EasyChem Pro, Systea Analytical Technology, Anagni, Italy). Ammonium-N and NO3-N were determined by extracting 5 g of soil with 25 mL of 2 M KCl and quantified by the modified indophenol blue technique (Sims et al., 1995) using a micro- plate spectrophotometer at 650 nm (Multiskan GO, Thermo Fisher Sci- entific, Waltham, MA). Olsen P was determined by extracting 2.5 g of soil with 25 mL of 0.5 M NaHCO3 (Olsen et al., 1954). Concentrations were quantified by colorimetry with a discrete analyzer (EasyChem Pro, Systea Analytical Technology, Anagni, Italy). Biochar was supplied by Cool Planet Energy Systems, Inc. (Green- wood Village, CO), which markets biochar products under the brand names CoolTerra® and CoolFauna®. The biochar provided was derived from pinewood (Pinus spp.) and created using the company’s proprietary Engineered Biocarbon™ technology, which includes a front-end biomass pyrolysis (<650 ◦C) and a patented post-pyrolysis treatment step. The biochar had an ash content of 1.7% as determined by standard ASTM 1762 methodology. It was characterized by a hydrogen:carbon ratio of 0.28, surface area of 152 m2 g− 1 (ASTM D6556), volatile matter of 25.4% (ASTM 1762) and bulk density of 122 kg m− 3 (dry mass basis; ASTM D2865) (InnoTech Alberta Inc., Vegreville, AB). Biochar (3 g) was added to a 50 mL Erlenmeyer flask with 30 mL of Ultrapure Milli-Q® water (≤18.2 MΩ cm− 1), mixed (30 min at 180 rpm) and then allowed to settle for 60 min at room temperature. The biochar pH and EC were determined in a 10:1 (water:biochar) slurry. Filtered biochar extracts (<5 µm, Q2 Fisherbrand™ Paper Circles) were analyzed for water- extractable NH4-N, NO3-N and PO4-P concentrations by colorimetry. Finely ground biochar samples (<0.15 mm) were used to determine TC, TN and TP concentrations, as previously described. Selected soil and biochar chemical properties are presented in Table 1. 2.2. Scanning electron microscopy (SEM) analysis The surface morphology of raw biochar was visualized using a Hitachi S-3400 N scanning electron microscope operated at 5.0 kV. Portions of biochar incubated in Raymond or Lethbridge soils were also analyzed using SEM. Briefly, a composite (n = 3) soil-biochar mixture was suspended in Ultrapure Milli-Q® water (10:1; water:soil) and shaken on a reciprocating shaker for 30 min. The resulting suspension was poured over a 0.210 mm sieve under a continuous flow of Ultrapure Milli-Q® water to ensure particle separation. All the material remaining on the screen (i.e., sand + biochar) was oven-dried at 40 ◦C for 24 h. Biochar particles > 2 mm were then hand-picked from the mixture with forceps and prepared for SEM analysis. Before SEM imaging, biochar particles were sputter-coated with gold to improve sample conductivity. C.M. Romero et al. Catena 198 (2021) 105046 3 2.3. Incubation experiment and soil column preparation An 86 d incubation experiment was conducted to investigate biochar effects on cumulative GHG emissions and soil nutrient availability in the presence or absence of NP-fertilizer. Treatments simulated 0, 0.15, 0.5, 3, 10 and 20 Mg biochar ha− 1 applied alone or with 150 kg N ha− 1 and 50 kg P ha− 1 NP-fertilizer. A total of twelve biochar-fertilizer combi- nations were applied to Raymond or Lethbridge soils. For each soil column, 0, 2.25, 7.5, 45, 150 or 300 mg of biochar (i.e., simulating 0, 0.15, 0.5, 3, 10 and 20 Mg ha− 1, respectively) were thoroughly mixed with 15 g of soil (dry weight equivalent) in a 50 mL glass beaker considering a soil depth of 10 cm and a bulk density of 1.0 g cm− 3. Urea- N (4.93 mg) and KH2PO4 (1.5 mg) were applied evenly to soils receiving NP fertilization to attain the aforementioned rates. Treatments were processed in a randomized complete block design with three replicates. Soil columns were constructed using 30 mL syringes following the protocol outlined in Campbell et al. (1993). The bottom of each syringe was lined with a Whatman glass microfiber filter (Grade GF/A, 1.6 μm) to prevent soil transport through the syringe tip, and the syringe tip was capped until sampling to prevent draining of water. Treated soil samples were mixed with laboratory-grade VWR Analytical BDH® Ottawa sand (15 g dry weight equivalent) and packed into columns to attain a bulk density of ~1.2 Mg m− 3. This approach produced a homogeneous mixture and prevented particle-size segregation throughout the experi- ment (Campbell et al., 1993). A glass-wool pad was placed over the contents of each column to prevent soil from dispersing when water was added (Fig. 1a). Ultrapure Milli-Q® water was added to each soil mixture to bring it to 60% of water holding capacity (WHC), pre- determined for each treatment. The tops of the columns were covered with perforated Parafilm® M to prevent anaerobic conditions from forming and reduce evaporation. The WHC was maintained at 60% by adding water (Ultrapure Milli-Q®) once weekly. When not being sampled, soil columns were kept at 21 ◦C. As a reference, soil columns (n = 3) were packed with 30 g of Ottawa sand and held at equivalent soil moisture levels. 2.4. Gas sampling and analysis Soil columns were used for CO2, N2O and CH4 measurements over a total of 20 sampling events (i.e., 0, 1, 3, 7, 8, 13, 14, 20, 21, 28, 29, 35, 40, 42, 47, 56, 71, 74, 79 and 86 d). Prior to gas sampling, the Parafilm® M cover was removed from the soil columns. Columns were placed in 1 L mason jars and sealed with lids equipped with a rubber septum at ambient air. Immediately after capping, 10 mL of gas was extracted from reference soil columns, i.e., sand-only, which was used to represent initial gas concentrations at time zero. Treated soil columns were incubated for 2 h and sampled to determine final gas concentrations. Gas samples were then immediately injected into pre-evacuated 5.8 mL vials. Samples were analyzed for CO2, N2O and CH4 concentrations using a gas chromatograph (Varian 3800, Varian Instruments, Walnut Creek, CA) equipped with thermal conductivity (TCD), flame ionization (FID) and electron capture detectors (ECD) (Fig. 1b). The injector and column temperatures were kept at 55 ◦C. The carrier was P10 gas (10% methane, balance argon) for the ECD and helium for the TCD and FID. The channel was maintained at a static pressure of 150 kPa. Gas concentrations were used to calculate fluxes (expressed as mg C for CO2 and CH4, and mg N for N2O, per kg of dry soil per unit time) using the (i) linear slope between time zero and two-hour gas concen- trations, (ii) dry mass of the soil in the column (excluding the Ottawa sand), (iii) jar headspace (adjusted for the soil column), (iv) standard atmospheric pressure, (v) air temperature and (vi) ideal gas law. All fluxes were extrapolated to daily averages. From preliminary trials, a two-hour incubation time consistently captured linear increases/decreases in gas concentrations. Therefore, only two samples were used to calculate our fluxes. Quadratic slopes of gas concentrations over time, often observed when measuring surface- to-atmosphere gas fluxes in the field, are the result of lateral diffusion of gases out of a static chamber into the soil, or ambient gases into a Table 1 Selected chemical properties of Raymond and Lethbridge (0–15 cm) soils and the applied pinewood biochar. Means are presented with standard errors. pH* EC* Total N Total C Total P NH4-N + NO3-N** PO4-P*** dS m− 1 g kg− 1 mg kg− 1 Raymond Soil 6.62 ± 0.09 0.09 ± 0.00 1.48 ± 0.02 14.44 ± 0.20 0.34 ± 0.02 6.67 ± 0.10 32.72 ± 1.71 Lethbridge Soil 7.86 ± 0.03 0.12 ± 0.00 1.83 ± 0.03 19.51 ± 0.30 0.67 ± 0.01 7.53 ± 0.09 35.23 ± 7.37 Biochar 7.17 ± 0.02 0.29 ± 0.01 1.58 ± 0.03 686.40 ± 2.50 0.15 ± 0.01 nd nd nd = below detection limit. * Measured in a 2:1 (soil) or 10:1 (biochar) water:sample slurry. ** 2 M KCl (soil) or water-extractable (biochar) NH4-N + NO3-N. *** 0.5 M NaHCO3 (soil) or water-extractable (biochar) PO4-P. Fig 1. Schematic representation of (a) soil column leaching experiment and (b) soil incubation and gas sampling. This figure was created using the website https: //app.biorender.com. C.M. Romero et al. https://app.biorender.com https://app.biorender.com Catena 198 (2021) 105046 4 static chamber contaminating the headspace (Creelman et al., 2013). This is not a factor in our study where soil columns are incubated in jars. Cumulative GHG emissions over the entire 86 d incubation period were calculated by linearly interpolating between measurements and summing daily fluxes to estimate total emissions, expressed as CO2-C, N2O-N and CH4-C mg kg− 1 soil. 2.5. Soil and leachate analyses Soil columns were retrieved at 7, 13, 20, 28, 40 and 71 d of incu- bation for NH4-N + NO3-N and PO4-P analyses. Leachates were obtained by adding 40 mL of 0.01 M CaCl2 in four increments of 10 mL each. Immediately after leaching, the bottom of each syringe was recapped and the top of each column covered with perforated Parafilm® M. Ex- tracts collected on each sampling day were stored at − 19 ◦C and analyzed within a week for water-extractable NH4-N + NO3-N and water-extractable PO4-P as described above (see Section 2.1). Available N (AN, mg kg− 1) and available P (AP, mg kg− 1) were indexed as the cumulative amount of water-extractable NH4-N + NO3-N or water- extractable PO4-P released over time, respectively. Final soil samples (86 d) were destructively analyzed for pH and EC, as well as water- extractable organic C [(WEOC); mg C kg− 1] and water-extractable N [(WEN); mg N kg− 1], quantified in syringe-filtered 15 mL aliquots (<0.45 µm) using a TC and TN combustion analyzer (TOC-VCSH and TNM-1 Shimadzu Corp., Kyoto, Japan) following Chantigny et al. (1999). 2.6. Statistical analysis Final soil properties, cumulative GHG emissions and available soil nutrient variables were analyzed using generalized linear mixed models (PROC GLIMMIX in SAS v. 9.4, SAS Institute Inc., Cary, NC). The models were mixed due to the inclusion of soil (S), fertilizer (F), biochar (B) and their two- and three-way interactions as fixed factors, and block as a random factor. To model the response variables, the log-normal (pH, CO2), gamma (WEOC, WEN, C/N, N2O and AN), Gaussian (EC, CH4) or inverse Gaussian (AP) distributions were selected based on the model fit statistics, i.e., the Bayesian Information Criterion. Post-hoc comparisons between treatment and interaction levels were conducted using Bonferroni-adjusted p-values; 95% confidence intervals (α = 0.05) were used to report differences between treatments. A multivariate redun- dancy analysis (RDA) was conducted to identify relationships between treatment variables (X) and cumulative soil nutrients and GHG emis- sions data (Y). An RDA is a multivariate multiple regression, followed by Fig. 2. Scanning electron microscopy (SEM) images of raw (a, b) and soil-aged pinewood biochar recovered from incubated Raymond (c, d) or Lethbridge (e, f) surface (0–15 cm) Chernozems. C.M. Romero et al. Catena 198 (2021) 105046 5 a principal component analysis (PCA) of fitted matrix values. The dec- orana function in the vegan package of R v. 3.4 (R Core Team, 2019) was used to confirm the global hypothesis of the linear relationship between X and Y (Oksanen et al., 2018). Figures were created using SigmaPlot v. 13.0 (Systat Software, San Jose, CA) and the vegan3d package of R v. 3.4. 3. Results and discussion 3.1. SEM characterization of biochar SEM imaging showed that the surface of raw biochar was rough and porous (Fig. 2a, b) and mainly comprised of particles with >10 µm macropores. A similar structure was observed for soil-aged portions of biochar extracted from Raymond (Fig. 2c, d) or Lethbridge (Fig. 2e, f); biochar recalcitrance was well-preserved despite mineral weathering (Dong et al., 2017). Furthermore, soil-aged biochar particles were coated with organo-mineral phases. This layer was presumably enriched by biochar-bound nutrients (Hagemann et al., 2017). 3.2. Soil properties 3.2.1. Soil pH and electrical conductivity Initial soil pH was near neutral in Raymond or slightly alkaline in Lethbridge (Table 1). After 86 d, soil pH ranged from 6.3 to 7.8 in Raymond, and from 7.1 to 7.9 in Lethbridge. The application of biochar with or without NP-fertilizer did not affect (p = 0.763) soil pH relative to the unamended control (Table 2). Final soil EC estimates ranged from 0.22 to 0.32 dS m− 1 in Raymond, and from 0.24 to 0.40 dS m− 1 in Lethbridge. The application of biochar or NP-fertilizer did not affect soil EC, i.e., p = 0.300 and p = 0.274, respectively (Table 2). However, the fertilizer × biochar interaction was significant (p = 0.015) after 86 d of incubation (Table 2). Increased soil pH following biochar amendment has been previously documented within acidic, highly weathered soils (Glaser et al., 2002; Major et al., 2010). Because most biochar is alkaline, its liming effect is now recognized as a management alternative for nutrient-impoverished fields with high Al+3 levels (El-Naggar et al., 2019a). In our study, the soil pH of neither Raymond nor Lethbridge changed due to biochar addition, possibly because the observed Chernozems were strongly buffered with high-activity clays (Brady and Weil, 2002) and exhibited an initial soil pH that was similar to the pH of the applied biochar, in agreement with Ahmed and Schoenau (2015). By contrast, Chathurika et al. (2016) reported a significant effect on soil pH when biochar (pH 9.7) was incorporated at 15–30 Mg ha− 1 to an incubated Manitoba Chernozem (pH 7.6). Our observation that biochar did not change the non-saline status of Raymond and Lethbridge was consistent with findings by Chathurika et al. (2016), who amended soil with pyrolyzed Manitoba Maple (Acer negundo L.) woodchips. Biochar has the potential to influence soil salinity, mainly through its ash content (El-Naggar et al., 2018). Nevertheless, changes in soil EC are biochar and soil dependent and range from little to no increase (Romero et al., 2021) to increases greater than 30% (El-Naggar et al., 2018). Kloss et al. (2014) found that soil EC increased after the application of biochar at 90 Mg ha− 1 to an Austrian Planosol. The effect varied with feedstock type, with wheat straw yielding higher soil EC (0.19 dS m− 1) relative to mixed woodchips (0.12 dS m− 1) or vineyard clippings (0.10 dS m− 1) (Kloss et al., 2014). By contrast, Soinne et al. (2014) observed a decrease in soil EC with biochar addition at 15 Mg ha− 1 to fine- and coarse-textured soils; biochar acted as a sorptive surface rather than a source of exchangeable ions (Soinne et al., 2014). Furthermore, it has been found that salt leaching is reduced by up to 18% in soil columns packed with corn-derived biochar at 45 Mg ha− 1 (Kanthle et al., 2016, 2018). The response was consistent among different soil types (e.g., Vertisols, Inceptisols) and land uses (e.g., forest, agricultural), and was attributed to higher water retention under bio- char treatments (Kanthle et al., 2016). 3.2.2. Water-extractable organic C and N concentrations The pool size of WEOC and WEN was determined at the end of the incubation period. After 86 d, the concentration of WEOC was nearly two-fold higher (p < 0.001) in Raymond (83.13 mg kg− 1) than in Lethbridge (46.47 mg kg− 1). The concentration of WEOC in Raymond and Lethbridge did not differ for either rate of biochar or NP-fertilizer relative to the unamended control (Table 2; Fig. 3a, b). The concentra- tion of WEN was affected by soil (p = 0.003) and the soil × fertilizer interaction (p = 0.016). Similar to WEOC, adding biochar up to 20 Mg ha− 1 did not affect (p = 0.206) the amount of WEN in Raymond (4.38–9.98 mg kg− 1) and Lethbridge (3.61–15.11 mg kg− 1) (Table 2; Fig. 3c, d). The C/N ratio of water-extractable OM ranged from 4.4 to 31.1 in Raymond, and from 2.7 to 11.9 in Lethbridge (Fig. 3e, f). Its distribution was affected by soil (p < 0.001), biochar (p = 0.006) and the soil × fertilizer interaction (p = 0.005) (Table 2). The application of biochar yielded similar WEOC concentrations, even though TC contents increased linearly with biochar rates >3 Mg ha− 1 in Raymond and Lethbridge (Fig. S1). Arable soils receiving considerable amounts of relatively labile OM (e.g., manure, straw or compost) will generally increase their WEOC pool size following the accumulation of TC in surface soil (Miller et al., 2018). Nevertheless, some studies have reported that increasing inputs of chemically-inert C such as biochar do not necessarily result in higher soil OM Table 2 Statistical significance (p-values) associated with the main effects: soil texture (sandy clay loam, clayey), biochar rate (0–20 Mg ha− 1) and NP-fertilization (-NP, +NP), and interactions as calculated by SAS PROC GLIMMIX for generalized linear mixed models (SAS v. 9.4). Variable Distribution Random residual group p-values of F-tests Soil (S) Fertilizer (F) Biochar (B) F × B S × B S × F S × F × B Soil properties pH log-normal Soil × Fertilizer <0.001 0.321 0.763 EC* Gaussian Fertilizer × Biochar <0.001 0.274 0.300 0.015 WEOC gamma Soil × Fertilizer <0.001 0.260 0.281 WEN gamma Soil 0.003 0.079 0.206 0.016 C/N gamma Soil × Fertilizer <0.001 0.729 0.006 0.005 Cumulative greenhouse (GHG) emissions N2O gamma Fertilizer 0.637 0.001 0.435 CO2 log-normal Biochar <0.001 0.632 0.006 0.028 0.001 CH4 Gaussian Soil <0.001 0.435 0.799 0.024 Available nutrients AP inv. Gaussian Soil × Fertilizer <0.001 <0.001 0.366 0.007 0.498 <0.001 0.024 AN gamma Soil × Fertilizer × Biochar 0.814 <0.001 0.554 <0.001 0.456 0.008 * EC (electrical conductivity), WEOC (water-extractable organic carbon), WEN (water-extractable nitrogen), C/N (water-extractable C/N ratio), AP (available phosphorus), AN (available nitrogen). C.M. Romero et al. Catena 198 (2021) 105046 6 concentrations in the soluble phase (Demisie et al., 2014); biochar may be a minor contributor of water-extractable OM over short- (Hernandez- Soriano et al., 2016), mid- (Kuzyakov et al., 2009) and long-term time intervals (Dong et al., 2019). The application of biochar yielded similar WEN concentrations, regardless of whether or not NP-fertilizer was added. Given that our biochar was characterized by negligible quantities of NH4-N and NO3-N, its application to Raymond or Lethbridge did not provide any additional labile-N. Our results agree with observations by Zavalloni et al. (2011) that soil soluble N and NO3-N pools are not affected by the presence of biochar. Estimates of water-extractable C/N ratio were in accordance with the literature (e.g., Toosi et al., 2012; Romero et al., 2017). The lower water- extractable C/N ratio of Lethbridge (~6.0) than Raymond (~12.0) appear to reflect a larger proportion of microbially-processed OM (Romero et al., 2017). The greater contribution of microbial by-products in fine- compared with coarse-textured soils has been attributed to the stabilization of N-rich compounds in organo-mineral complexes (Samson et al., 2020). Similar to previous studies, the application of biochar to soil, a material with a considerably higher total C/N ratio (>400) than Raymond or Lethbridge, did not result in a measurable change of water-extractable C/N ratios (Table S1). Romero et al. (2021) reported that biochar (3 Mg ha− 1) and biochar-manure (120 Mg ha− 1) affected water-extractable C/N ratios of incubated Chernozems to a minor degree. Given that the biodegradability of biochar is minimal, its incorporation to soil may have a neutral effect on the microbial pro- cessing of total OM and its subsequent mineralization in the soluble phase (Zavalloni et al., 2011; Hernandez-Soriano et al., 2016; Yanagi et al., 2016), even when co-applied with NP-fertilizer (Zhang et al., 2017b). Alternatively, wood-derived biochar may immobilize lower amounts of NO3-N and NH4-N relative to more labile-C substrates, pre- cluding shifts towards higher C/N ratios (Zavalloni et al., 2011). 3.3. Greenhouse gas emissions 3.3.1. Carbon dioxide Soils from Raymond (Fig. 4a, e) and Lethbridge (Fig. 4b, f) exhibited similar CO2-C flux trends, regardless of NP-fertilizer addition. Initially, Raymond CO2-C fluxes increased with time, irrespective of biochar application rate, and then sharply decreased to <2.75 mg CO2-C kg− 1 soil h− 1 after 3 d. Thereafter, Raymond CO2-C fluxes remained almost stable until the end of the study in a manner similar to those observed with Lethbridge. Cumulative CO2-C emissions ranged from 1313.31 to 2644.35 mg kg− 1 in Raymond, and from 823.89 to 2060.89 mg kg− 1 in Lethbridge (Fig. 4a, b, e, f), and were affected by soil (p < 0.001) and biochar (p = 0.006), and the fertilizer × biochar (p = 0.028) and soil × fertilizer (p = 0.001) interactions (Table 2). Biochar-induced reductions on soil CO2-C fluxes have been widely-documented under laboratory and field conditions (Al-Wabel et al., 2018; El-Naggar et al., 2019b). For example, Tian et al. (2019) reported reduced soil respiration upon amending a clayey soil with straw-derived biochar relative to straw- only. Applying biochar induced a microbial community shift towards soil bacteria with strong lignin affinity (e.g., actinobacteria), thereby reducing OM mineralization by limiting enzyme activity (e.g., β-gluco- sidase, dehydrogenase) over labile-C forms (Tian et al., 2019). Never- theless, the extent to which biochar inhibits the rate of CO2-C emission varies depending on the type of biochar feedstock (e.g., wood vs. straw), initial soil properties (e.g., texture) and N-fertilizer inputs (Kammann et al., 2017; El-Naggar et al., 2019b). El-Naggar et al. (2018) reported that the application of biochar at 30 Mg ha− 1 did not affect cumulative CO2-C emissions in a sandy loam soil with 1.90% OM, but induced a positive priming effect when applied to a sandy soil with 0.57% OM. In Alberta, Pokharel et al. (2018) reported that biochar did not affect CO2- C fluxes from a grassland Chernozem, but effectively reduced cumula- tive CO2-C emissions within a forest Luvisol. Sui et al. (2016) observed that co-application of straw-derived biochar (14.8 Mg ha− 1) with urea-N (210 kg N ha− 1) markedly increased cumulative CO2-C emissions rela- tive to a biochar-only treated soil. This presumably resulted from the surplus of mineral-N enhancing the mineralization of labile-C within biochar. Using a different feedstock, Li et al. (2017) reported higher CO2-C fluxes upon amending an incubated soil with increasing amounts of apple-branch (Malus pumila Mill.) biochar, regardless of N-fertiliza- tion. In our study, amending soil with biochar (at different rates) affected cumulative CO2-C emissions (p < 0.001), but not in soil Fig. 3. Box and whisker plots of (a, b) water-extractable organic carbon (WEOC, mg kg− 1), (c, d) water- extractable nitrogen (WEN, mg kg− 1) and (e, f) water-extractable C/N ratio of Raymond and Lethbridge soils, respec- tively. Box plots show the 25 and 75 percentile ranges; solid and dashed inbox lines represent the median and mean, respectively; whiskers delimit the non-outlier ranges; and the symbol (x) represents outlier values. For each var- iable, means followed by a common letter are not significantly different (Bonferroni Grouping at p < 0.05 for the soil × fertilizer interaction). C.M. Romero et al. Catena 198 (2021) 105046 7 receiving NP-fertilizer (p = 0.345). Similarly, NP-fertilization decreased (p = 0.009) cumulative CO2-C emissions only at 0.15 Mg ha− 1 and increased (p = 0.015) cumulative CO2-C emissions in the 10 Mg ha− 1 treatment. NP-fertilizer application reduced cumulative CO2-C emis- sions in Lethbridge (p = 0.015), but not in Raymond, where the results were suggestive but inconclusive for higher cumulative CO2-C emissions with the addition of urea-N (p = 0.080) (data not shown). This could be attributed to the fact that WEN’s response to N-fertilizer was greater in Raymond than Lethbridge (Fig. 3c, d), presumably increasing OM mineralization by narrowing water-extractable C/N ratios (Toosi et al., 2012). 3.3.2. Nitrous oxide Nitrous oxide fluxes were similar between Raymond and Lethbridge (<0.02 mg N2O-N kg− 1 soil h− 1), with modest N2O-N peak fluxes (<0.04 mg N2O-N kg− 1 soil h− 1) occurring after leaching of soil columns (8, 14, 21, 29, 42 and 74 d). Cumulative N2O-N emissions ranged from 1.12 to 16.94 mg kg− 1 in Raymond (Fig. 4c, g), and from 2.75 to 15.03 mg kg− 1 in Lethbridge (Fig. 4d, h), and were only affected by fertilizer (p = 0.001) (Table 2). Our results suggest that biochar did not reduce cumulative N2O-N emissions, even when co-applied with urea-N. This is in contrast with Wu et al. (2013) and Pokharel et al. (2018), who re- ported significant reductions in cumulative N2O-N emissions (i.e., 66% and 15%, respectively) when biochar was incubated with Chernozems from Alberta. In Saskatchewan, Hangs et al. (2016) found that net N2O- N emissions from Black Chernozems (0–15 cm) were reduced by 66% and 59% when shrub willow (Salix spp.) biochar (20 Mg C ha− 1) was applied with or without urea-N, respectively. A review by Nguyen et al. Fig. 4. Temporal dynamics of CO2-C and N2O-N fluxes (mg kg− 1 soil h− 1) and cumulative CO2-C and N2O-N emissions (mg kg− 1 soil) from surface (0–15 cm) Chernozems of (a), (c), (e) and (g) sandy clay loam (Raymond), and (b), (d), (f) and (h) clayey (Lethbridge) textures. Note: white- and gray-colored graphs depict un- fertilized and fertilized soil columns, respectively. Asterisks depict soil leaching events. Differences in cumulative N2O emissions among biochar rates are numeric but not statistically significant p < 0.05. (For interpretation of the references to colour in this figure legend, the reader is referred to the web version of this article.) C.M. Romero et al. Catena 198 (2021) 105046 8 (2017) highlighted four general pathways whereby biochar could reduce soil N2O-N fluxes, namely (i) binding N2O-N with metal ions, (ii) reducing the availability of labile-C and inorganic-N substrates, (iii) improving soil aeration and (iv) pH-induced shifts in the abundance/ activity of denitrifying bacteria. Similar to CO2-C emissions, N2O-N flux responses to biochar are also affected by soil texture and biochar feed- stock type. Zheng et al. (2012) reported significant N2O-N flux re- ductions from a biochar-amended fine-loamy Mollisol fertilized with 100 kg N ha− 1, but did not observe the same response within a fine- smectitic Mollisol. In Alberta, Cheng et al. (2012) reported that co- application of wheat-straw biochar with fertilizer-15N did not influ- ence cumulative N2O-N emissions from an incubated Chernozem. The authors attributed this response to a lack of labile-C supply for microbial activity; biochar had a negligible effect on gross N transformation rates relative to its uncharred feedstock (Cheng et al., 2012). Likewise, Romero et al. (2021) reported similar N2O-N fluxes after amending Raymond (<12 mg N2O-N kg− 1 soil) and Lethbridge (<10 mg N2O-N kg− 1 soil) with pinewood biochar, cattle manure or a biochar-manure mixture. This is consistent with the present findings, and further sug- gests that co-applying biochar with an external inorganic- or organic-N source may have a similar effect on cumulative N2O-N emissions. 3.3.3. Methane Methane flux rates were low during the incubation period, never exceeding 0.004 or 0.030 mg CH4-C kg− 1 h− 1 in Raymond and Leth- bridge, respectively (data not shown). Cumulative CH4-C emissions ranged from − 8.86 to − 0.33 mg kg− 1 and from − 4.72 to 8.80 mg kg− 1 in Raymond and Lethbridge, respectively, and were affected by soil (p < 0.001) and the soil × fertilizer interaction (p = 0.024) (Tables 2 and 3). Small fluxes in our study are in agreement with previous reports (Zhang et al., 2017a; Htun et al., 2017) that the release of CH4-C by rainfed croplands is limited (Conrad, 1996). It has been suggested that biochar amendment might increase soil aeration, presumably reducing CH4-C production and/or augmenting CH4-C oxidation (Karhu et al., 2011). However, our results indicate that biochar did not affect cumulative CH4-C emissions. Similar to our work, Wu et al. (2013) amended an Orthic Black Chernozem from southeastern Alberta with biochar at 10 and 25 Mg ha− 1 and did not observe a decrease in CH4-C emissions when compared to an unamended control. In Saskatchewan, Hangs et al. (2016) incorporated shrub willow biochar into Brown or Black Cher- nozems and concluded that biochar increased the uptake of CH4-C compared to an unamended control by 90% and 139%, respectively, but this response was negated if urea-N was co-applied at 100 kg N ha− 1. A recent meta-analysis by Jeffery et al. (2016) indicated that the addition of biochar to non-flooded systems might decrease the CH4 sink capacity of neutral or slightly alkaline soils, but only when co-applied with N- fertilizer at <120 kg N ha− 1. This can be attributed to a shift from methanotrophy to nitrification induced by the accumulation of NH4 + following urea-N hydrolysis (Acton and Baggs, 2011). The negative ef- fect of N-fertilizer on CH4-C oxidation has been reported in a wide range of habitats, including forest, grassland and agricultural ecosystems (Bodelier and Laanbroek, 2004). In our study, an overall reduction in net CH4-C uptake due to urea-N application was evident in Raymond, where average CH4-C gains were reduced by ~23% relative to biochar-only treated soil (p < 0.001). This response was not observed for Leth- bridge (p = 0.405), implying that soil texture could regulate the turnover rate of urea-N (Galantini et al., 2004) and the activity of methanotrophs (Boeckx et al., 1997). 3.4. Available nitrogen and phosphorus pools The pool size of AN and AP was determined as the cumulative amount (based on six sampling events) of water-extractable N and P leached from Raymond and Lethbridge. After 86 d, AN concentrations were similar between soils and ranged from 85.68 to 167.28 mg kg− 1 within the –NP treatment and from 353.41 to 1312.86 mg kg− 1 within the +NP treatment (Table 4). The effect of fertilizer (p < 0.001), and the fertilizer × biochar (p < 0.001) and soil × fertilizer (p = 0.008) in- teractions were significant for AN (Table 2). Available P was affected by soil (p < 0.001) and fertilizer (p < 0.001), and the fertilizer × biochar (p Table 3 Cumulative net CH4-C uptake (n = 3) within surface (0–15 cm) Chernozems of sandy clay loam (Raymond) and clayey (Lethbridge) textures 86 d after amendment with pinewood biochar as affected by biochar rate (0–20 Mg ha− 1) within columns receiving 0 kg N ha− 1 and 0 kg P ha− 1 (-NP) or 150 kg N ha− 1 and 50 kg P ha− 1 (+NP). Means are presented with standard errors. Biochar rate (Mg ha− 1) Raymond (mg kg− 1 soil) Lethbridge (mg kg− 1 soil) − NP +NP − NP +NP 0 − 6.83 ± 1.04* − 5.38 ± 1.11 − 0.31 ± 1.03 − 2.16 ± 0.70 0.15 − 6.21 ± 0.68 − 4.52 ± 1.16 − 0.57 ± 1.81 − 1.14 ± 1.14 0.5 − 6.85 ± 1.20 − 3.63 ± 1.93 0.14 ± 0.48 − 2.37 ± 0.58 3 − 6.21 ± 1.76 − 5.93 ± 1.29 − 0.02 ± 4.41 − 2.06 ± 1.20 10 − 6.64 ± 1.79 − 4.44 ± 0.43 − 2.67 ± 0.23 − 1.91 ± 0.48 20 − 5.55 ± 0.91 − 5.50 ± 1.32 − 2.86 ± 0.32 − 1.02 ± 1.21 * Within soils, differences are numeric but not statistically significant at p < 0.05. Table 4 Cumulative available N (AN) distribution after 86 d of incubation within surface (0–15 cm) Chernozems of sandy clay loam (Raymond) and clayey (Lethbridge) textures as affected by biochar rate (0–20 Mg ha− 1) within columns receiving 0 kg N ha− 1 and 0 kg P ha− 1 (-NP) or 150 kg N ha− 1 and 50 kg P ha− 1 (+NP). Means are presented with standard errors. Biochar rate (Mg ha− 1) Raymond (mg kg− 1 soil) Lethbridge (mg kg− 1 soil) − NP +NP − NP +NP 0 138.42 ± 17.14b* 822.85 ± 87.01a 141.39 ± 12.13b 879.29 ± 280.79a 0.15 145.98 ± 1.83b 860.58 ± 30.10a 107.05 ± 11.12b 928.92 ± 99.76a 0.5 130.85 ± 14.88b 878.31 ± 88.68a 123.18 ± 5.11b 1038.78 ± 76.40a 3 141.41 ± 11.69b 1064.64 ± 133.46a 126.34 ± 13.45b 1002.60 ± 13.79a 10 116.34 ± 15.15b 973.83 ± 72.58a 98.30 ± 11.20b 1072.49 ± 37.91a 20 121.45 ± 5.24b 1005.57 ± 57.21a 123.17 ± 20.06b 1141.47 ± 86.37a * Within a row, means followed by a common letter are not significantly different (Bonferroni Grouping at p < 0.05 for the biochar × fertilizer interaction). C.M. Romero et al. Catena 198 (2021) 105046 9 = 0.007), soil × fertilizer (p < 0.001) and soil × fertilizer × biochar (p = 0.024) interactions. Available P vs. biochar rate conformed to an exponential rise within fertilized-NP treated columns (Fig. 5). Biochar- derived AP gains proceeded more rapidly and plateaued at 3 Mg ha− 1 for Raymond while it increased with all increases in biochar rate in Lethbridge (Fig. 5). Available N was mostly unaffected by the addition of biochar to Raymond or Lethbridge. Amending soils with biochar has been shown to temporarily reduce NO3-N and NH4-N availability through microbial (Bruun et al., 2012) and abiotic pathways (Lehmann and Joseph, 2009). The magnitude of this effect depends mainly on biochar feedstock type and pyrolysis temperature (Ippolito et al., 2012; Nguyen et al., 2017). Wood-derived biochars typically induce smaller reductions in AN rela- tive to other sources of pyrolyzed OM enriched by labile-C pools (Nguyen et al., 2017). Our biochar was characterized by a relatively low volatile matter content (25.4%), indicating a low level of water- extractable OM availability for microbial N uptake (Nelson et al., 2011). The absence of significant effects of biochar rate on AN may also be attributed to the negligible quantities of NH4-N and NO3-N extracted from the raw material. Biochar-induced benefits on AN were only observed following the co-application of biochar with urea-N at 150 kg N ha− 1. Biochar, combined with urea-N, may improve soil NO3-N availability through increased nitrification (Nguyen et al., 2017). Rela- tive to an N-fertilized control, Song et al. (2014) reported increased rates of ammonia oxidation and a higher abundance of nitrifiers within a fertilized [(NH4)2SO4] alkaline soil (pH 8.0) amended with biochar derived from cotton (Gossypium hirsutum L.) stalks. Optimal pH of local soil microsites (Ball et al., 2010), elimination of inhibitory compounds (DeLuca et al., 2006), increased soil moisture (DeLuca et al., 2009), physical protection of microbial populations (Quilliam et al., 2013; DeLuca et al., 2009) and higher substrate availability in the charosphere (Yu et al., 2019) are some of the potential mechanisms that may favor nitrification in biochar-amended soils. However, it is important to note that the effect of biochar and urea-N co-application on AN was less pronounced than that of AP to biochar and KH2PO4 treatment. Our re- sults agree with observations by Song et al. (2014) that the potential for soil ammonia oxidation is not linearly related to the rate of biochar application. Pyrolysis of high-P enriched feedstocks may increase soil TP by converting organic P forms into phosphate minerals of varying biolog- ical and chemical reactivity (Sun et al., 2018). The application of bio- char may indirectly benefit soil TP availability by (i) pH-mediated reactions, (ii) stimulating P solubilizing microorganisms and (iii) coupling P sorption-desorption processes (Gul and Whalen, 2016; Sun et al., 2018). Biochar may also act as a direct P source, enhancing TP concentrations in surface soil (Atkinson et al., 2010; El-Naggar et al., 2018). In our study, AP was largely unaffected by the addition of biochar to Raymond or Lethbridge. It is likely that this can be partly attributed to minimal biochar-induced OM mineralization and enzymatic P activity because biochar amendment did not affect cumulative CO2-C emissions or water-extractable OM concentrations. Also, biochar application did not induce changes to soil pH, which may have inhibited the desorption of PO4-P from Fe+3 or Al+3 oxides (Dudas and Pawluk, 1969). This is supported by our observation that biochar only increased AP in Ray- mond or Lethbridge when co-applied with KH2PO4 at 50 kg P ha− 1. Sun et al. (2018) demonstrated that biochar precludes P-fertilizer from precipitating into secondary minerals (e.g., apatite), thereby increasing NaHCO3-P or NaOH-P pools in Chernozems (Chathurika et al., 2016). An alternative explanation is that biochar may delay KH2PO4 precipitation by sorbing soluble cations (e.g., Ca+2 or Mg+2). This response was pre- viously reported by Nelson et al. (2011) in an incubated Mollisol from Kansas that was amended with corn cob-derived biochar. Biochar may act as a “carrier of nutrients” for improved agronomic performance (Al-Wabel et al., 2018; El-Naggar et al., 2019a). Soil nu- trients retained by biochar are presumably delivered to plants at a much slower rate than those from manure, compost or sludge (El-Naggar et al., 2019a). Coating NP-fertilizer with biochar or “fertichars” is a promising approach to enhancing nutrient cycling in the rhizosphere (Kammann et al., 2017). Shi et al. (2020) reported that blending urea-N with bio- char and bentonite enhanced corn N use efficiency through increased plant N uptake and reduced NO3-N leaching in soil. The biochar-urea composite was shown to retain dissolved nutrients and maintain NO3 − / NH4 + ratios by binding N-fertilizer crystals to biochar/mineral surfaces. 3.5. Relationships between soil variables, GHG emissions and biochar application We performed an RDA to explain the relationships between treat- ment variables (X) that are known to affect the cycling of soil nutrients such as biochar application rate, NP-fertilization and soil texture, and response variables (Y) that are indicative of soil fertility and environ- mental performance, such as nutrient availability (AP, AN), water- extractable OM pools (WEOC, WEN) and cumulative GHG emissions. Fig. 5. Cumulative available P (PO4-P) distribution after 86 d of incubation within surface (0–15 cm) Chernozems of (a) clayey (Lethbridge) and (b) sandy clay loam (Raymond) textures as affected by biochar rate (0–20 Mg ha− 1) within columns receiving 0 kg N ha− 1 and 0 kg P ha− 1 (− NP) or 150 kg N ha− 1 and 50 kg P ha− 1 (+NP). C.M. Romero et al. Catena 198 (2021) 105046 10 The resulting partitioning of correlations indicated that the linear ad- ditive combination of X explained 56.43% of the total variation in the matrix of Y, with 36.49%, 10.83% and 9.11% of the total variance being explained by the first, second and third RDA axes, respectively (i.e., RDA1 to RDA3). The RDA results were represented in a correlation triplot (scaling = 2) (Fig. 6; Fig. S2). The angles between the vectors of biochar rate, cumulative N2O-N emissions and AN were small, indi- cating the positive correlations of these variables. A positive correlation was also observed between biochar rate, TC and total C/N within the upper right quadrant of the triplot. This response was anticipated given the fact that biochar is known to increase soil OM pools (Al-Wabel et al., 2018) by adding aromatic forms of C that are typically N-deprived (El- Naggar et al., 2019b; Liu et al., 2019). The centroid for Lethbridge and the variables TN, TP, EC, pH and cumulative CH4-C emissions were all located in the lower right quadrant of the triplot. Similarly, NP-fertilizer, WEN and AP were located in the same 3-D octant, indicating a strong dependence of these variables to the combined input of urea-N and KH2PO4. The centroid for Raymond was located in the upper left quadrant of the triplot and was correlated to WEOC and cumulative CO2- C emissions, indicating faster turnover rates of water-extractable OM within Raymond (i.e., sandy clay loam) than Lethbridge (i.e., clayey). 4. Conclusions The results of our study partially supported our hypothesis that amending soil with biochar would benefit the fertility status of the studied Chernozems. The application of biochar had no effect on cu- mulative CH4-C and N2O-N emissions and inorganic-N and -P avail- ability. Cumulative CO2-C emissions were different between soil textures and were increased or decreased non-linearly by biochar addition under − NP only. After 86 d, biochar amendment yielded similar WEOC concentrations, even though soil TC contents linearly increased with biochar rates >3 Mg ha− 1 in both Raymond and Leth- bridge. A longer time-scale might be required, even under controlled conditions, for biochar to affect OM mineralization and nutrient cycling in well-buffered Chernozems. However, co-applying biochar with NP- fertilizer provided a benefit to AP and AN without increasing soil pH, implying that soil-aged biochar particles favored plant-available P and N pools. The mechanisms driving the occurrence and subsequent release of biochar-bound NP-fertilizer warrants further investigation. This would require field studies to assess whether co-applications of biochar and NP- fertilizer are beneficial for the cycling of soil nutrients within intensively managed agroecosystems of western Canada. Declaration of Competing Interest The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper. Acknowledgments This work was funded by the University of Lethbridge research project “Assessment of the potential for adding biochar to beef cattle diets to reduce greenhouse gas emissions in agriculture” (AGGP2-037) granted to Dr. Erasmus Okine in partnership with the Agricultural Greenhouse Gases Program (AGGP) of Agriculture and Agri-Food Can- ada. We thank Jessica Stoeckli, Ryan Beck and Colten Gurr for their assistance in soil sampling and laboratory analysis. SEM analysis of biochar samples was performed by Grant Duke (AAFC - Lethbridge Research and Development Center). Biochar was provided by Cool Planet Energy Systems, Inc. (Greenwood Village, CO). We gratefully acknowledge the constructive input provided by two anonymous re- viewers and Catena’s Associate Editor. Appendix A. Supplementary material Supplementary data to this article can be found online at https://doi. org/10.1016/j.catena.2020.105046. References Acton, S.D., Baggs, E.M., 2011. Interactions between N application rate, CH4 oxidation and N2O production in soil. Biogeochemistry 103, 15–26. Al-Wabel, M.I., Hussain, Q., Usman, A.R., Ahmad, M., Abduljabbar, A., Sallam, A.S., Ok, Y.S., 2018. Impact of biochar properties on soil conditions and agricultural sustainability: a review. L. Degrad. Dev. 29, 2124–2161. Ahmed, H.P., Schoenau, J.J., 2015. Effects of biochar on yield, nutrient recovery, and soil properties in a canola (Brassica napus L)-wheat (Triticum aestivum L) rotation grown under controlled environmental conditions. Bioenerg. Res. 8, 1183–1196. Atkinson, C.J., Fitzgerald, J.D., Hipps, N.A., 2010. Potential mechanisms for achieving agricultural benefits from biochar application to temperate soils: a review. Plant Soil 337, 1–18. Ball, P.N., MacKenzie, M.D., DeLuca, T.H., Montana, W.E., 2010. Wildfire and charcoal enhance nitrification and ammonium-oxidizing bacterial abundance in dry montane forest soils. J. Environ. Qual. 39, 1243–1253. Bodelier, P.L., Laanbroek, H.J., 2004. Nitrogen as a regulatory factor of methane oxidation in soils and sediments. FEMS Microbiol. Ecol. 47, 265–277. Boeckx, P., Van Cleemput, O., Villaralvo, I., 1997. Methane oxidation in soils with different textures and land use. Nutr. Cycl. Agroecosystems 49, 91–95. Brady, N., Weil, R., 2002. The Nature and Properties of Soils. Pearson Education Inc, Upper Saddle River, NJ. Bruun, E.W., Ambus, P., Egsgaard, H., Hauggaard-Nielsen, H., 2012. Effects of slow and fast pyrolysis biochar on soil C and N turnover dynamics. Soil Biol. Biochem. 46, 73–79. Campbell, C.A., Ellert, B.H., Jame, Y.W., 1993. Nitrogen mineralization potential in soils. In: Carter, M.R. (Ed.), Soil Sampling and Methods of Analysis. Lewis Publishers, Boca Raton, FL, pp. 341–349. CanSIS, Canadian Soil Information Service, 2013. Soils of Alberta. http://sis.agr.gc.ca/ cansis/soils/ab/soils.html (accessed 02 May 2019). Case, S.D.C., McNamara, N.P., Reay, D.S., Whitaker, J., 2012. The effect of biochar addition on N2O and CO2 emissions from a sandy loam soil – the role of soil aeration. Soil Biol. Biochem. 51, 125–134. Fig. 6. Correlation triplot of soil properties, nutrient availability, cumulative greenhouse gas emissions and water-extractable organic matter from redun- dancy analysis (RDA) for soil texture [clayey (Lethbridge) and sandy clay loam (Raymond)] and management factors including biochar rate (0–20 Mg ha− 1) and NP-fertilization (− NP, +NP). Hollow circles represent observed values of the dependent/response variables. Blue arrows and associated blue text indicate the vectors of quantitative explanatory variables. Blue ellipses indicate the re- gions of 95% confidence for the locations of centroids associated with the categorical explanatory variable Raymond or Lethbridge soil. Total carbon, TC; total nitrogen, TN; total phosphorus, TP; electrical conductivity, EC; available nitrogen, AN; available phosphorus, AP; water-extractable organic carbon, WEOC; water-extractable nitrogen, WEN. Note: the locations of centroid labels were adjusted for readability. (For interpretation of the references to colour in this figure legend, the reader is referred to the web version of this article.) C.M. 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