This is the accepted version of the article: Andrés Pastor, Pilar; Rosell-Melé, Antoni; Colomer Ventura, Ferran; [et al.]. Belowground biota responses to maize biochar addition to the soil of a Mediter- ranean vineyard. DOI 10.1016/j.scitotenv.2019.01.101 This version is avaible at https://ddd.uab.cat/record/203557 under the terms of the license 1 BELOWGROUND BIOTA RESPONSES TO MAIZE BIOCHAR 2 ADDITION TO THE SOIL OF A MEDITERRANEAN VINEYARD 3 Pilar Andrés a,b*, Antoni Rosell-Melé b,f, Ferran Colomer-Venturaa, Karolien Denef c, M. 4 Francesca Cotrufo d, Miquel Riba a,e, Josep M. Alcañiz a,e a 5 CREAF. Edifici C, Campus UAB. 08193, Cerdanyola del Vallès. Barcelona, Spain. b 6 ICTA, Edifici ICTA-ICP, Carrer de les Columnes s/n, Campus UAB. 08193 Cerdanyola del 7 Vallès, Barcelona, Spain. c 8 Central Instrument Facility, Chemistry Department. Colorado State University. 1301 Center 9 Avenue. Campus Delivery 1872. Fort Collins. CO 80523-1872, USA d 10 Natural Resource Ecology Laboratory & Soil and Crop Sciences Department. Colorado State 11 University. 200 W. Lake. Campus Delivery 1499. Fort Collins. CO 80523-1499, USA e 12 Department of Animal Biology, Plant Biology and Ecology. Faculty of Sciences. Autonomous 13 University of Barcelona. Edifici C, Campus UAB. 08193, Cerdanyola del Vallès. Barcelona, 14 Spain. f 15 Institució Catalana de Recerca i Estudis Avançats (ICREA), Passeig Lluís Companys 23. 16 08010 Barcelona, Spain. 17 18 * Corresponding author: pilar.andres@uab.es 19 20 Abstract 21 Biochar is a high carbon material resulting from biomass pyrolysis that, when applied 22 to croplands, can increase soil carbon and soil water retention. Both effects are of critical 23 importance in semi-arid regions, where carbon decline and desertification are the main 24 drivers of soil degradation. Since most environmental services provided by soil are 25 mediated by belowground biota, effects of biochar on soil microbial and invertebrate 26 communities must be evaluated under field conditions before its agricultural application 27 can be recommended. We tested maize biochar for its mid-term effect on soil microbes 28 and micro-arthropods of a Mediterranean vineyard. We applied biochar to three field plots 1 29 with neutral sandy loam soils at a dose of 5 Mg ha-1. During two years, we monitored the 30 abundance of functional groups of soil micro-arthropods and estimated the biomass of 31 soil microbial groups. We also analyzed the δ13C value of microbial PLFA biomarkers to 32 determine biochar-C utilization by each microbial group taking advantage of the δ13C 33 natural abundance differences between the applied biochar and the soil. Biochar addition 34 significantly reduced soil microbial biomass but did not alter the functional microbial 35 diversity nor the abundance or biodiversity of soil micro-arthropods. The contribution of 36 biochar-C to the diet of most microbial groups was very low through the monitoring 37 period. However, two gram-negative bacterial groups increased their biochar-derived 38 carbon uptake under extreme soil dryness, which suggests that biochar-C might help soil 39 microbes to overcome the food shortage caused by drought. The decrease in microbial 40 biomass observed in our experiment and the concomitant decrease of SOM mineralization 41 could contribute to the carbon sequestration potential of Mediterranean soils after biochar 42 addition. 43 44 Keywords: biochar, Mediterranean soils, soil biota, soil microbial biomass, microbial 45 biochar utilization, PLFA. 46 47 1. Introduction 48 Counteracting soil carbon decline is a key priority for sustainable soil management in 49 the arid and semiarid areas of Europe, since there is evidence that soil degradation will 50 further progress as climate variability increases and extreme weather events become more 51 frequent (Montanarella, 2007). Biochar production and application to soil is promoted as 52 a way to increase the recalcitrant soil carbon pool while improving soil water-holding 53 capacity (Atkinson et al., 2010). Biochar is a by-product of the pyrolysis of biomass at 2 54 temperatures ranging from 350ºC to more than 800ºC in the absence (or at very low 55 concentration) of oxygen (Sohi et al., 2009). Agricultural lands produce high quantities 56 of organic residues that, when naturally returned to soil, are efficiently processed by the 57 underground food web. After 5 to 10 years, 80 to 90% of the C-biomass will have been 58 released back to the atmosphere as CO2. Pyrolysis of these residues leads to sequestration 59 of about 50% of their carbon into recalcitrant biochar-C with an estimated residence time 60 in soil of hundreds to thousands of years (Lehmann et al., 2006). For this reason, biochar 61 production and application to agricultural soils has been suggested as a potential strategy 62 to develop more sustainable agricultural systems while mitigating greenhouse gas 63 emissions (Roberts et al., 2009; Woolf et al., 2010). 64 Biochar is claimed to enhance crop yields (Atkinson et al., 2010; Spokas et al., 2012). 65 However, increases in crop production have only been proven for nutrient-poor acidic 66 and coarse and medium textured soils, while moderately fertile arable soils in temperate 67 regions rarely show significant yield increases after biochar application (Sorrenti et al., 68 2016; Agegnehu et al., 2017; Jeffery et al., 2017). Together with improvement in soil 69 physical conditions and nutrient status (Biederman and Harpole, 2013), the liming effect 70 of biochar is thought to be the main mechanism underlying yield increase in acidic soils 71 (Jeffery et al., 2011). In neutral to basic and light-textured soils under temperate and dry 72 climates, the agricultural benefits of biochar are more often attributable to the 73 improvement of soil water-holding capacity (Olmo et al., 2014; Baronti et al., 2014; 74 Genesio et al., 2015). 75 Most environmental services provided by soil, including agricultural fertility, carbon 76 sequestration and water cycle regulation are substantially mediated by the activity of a 77 highly diverse soil community of microbes and invertebrate animals (Lavelle et al., 2006). 78 In the surface horizon of temperate agricultural soils, microbial biomass is in the range of 3 79 400 to 880 mg C kg soil-1 (Dalal, 1998) and animals are present at densities of 106 m-2 for 80 nematodes, 105 m-2 for micro-arthropods and 104 m-2 for other invertebrates (Altieri et al., 81 1999). Soil multifunctionality and even plant biodiversity are closely dependent on soil 82 communities (Wagg et al., 2014) and the interaction between plants and belowground 83 organisms regulates primary production and plant health (Wardle et al., 2004). Under the 84 current scenario of climatic uncertainty, sustainable agricultural management must aim 85 to increase soil ecosystem resilience that is closely dependent on soil biodiversity and 86 belowground food web structure (Andrés et al., 2017). 87 Biochar can influence belowground communities through changes in soil albedo (Meyer 88 et al., 2012), soil chemistry and physical structure, moisture and aeration (Atkinson et al., 89 2010), nutrient availability, pH (McCormack et al., 2013) and toxicity (Hilber et al., 90 2017), and by providing a carbon source to the soil biota (Soong et al., 2017). Soil 91 porosity highly determines microbial abundance because a large proportion of the soil 92 bacteria live in micropores inside and around soil microaggregates that offer favorable 93 conditions of water and substrate availability and protection against predators (Rabbi et 94 al., 2016; Sessitsch et al., 2001). Thanks to its highly porous structure, biochar is 95 considered a good soil conditioner but its efficiency in improving soil porosity depends 96 on pore size distribution of the biochar particles that varies with pyrolysis conditions and 97 biomass feedstock (Downie et al., 2009). Both the type of pyrolysis and the chemistry of 98 the feedstock will also determine the recalcitrance of biochar to mineralization and the 99 amount of nutrients available to microbes that ultimately determines biochar stability and 100 the overall biochar-C sequestration capacity of the soil (Schlesinger and Andrews, 2000; 101 Thies and Rillig, 2009). 102 It is widely recognized that biochar degradation is both biotically and abiotically 103 mediated (Jones et al., 2011) and that its application to soil may have significant effects 4 104 on belowground processes. Therefore, consequences of biochar on the soil microbial 105 community must be evaluated before agricultural application of biochar can be 106 recommended (Pressler et al., 2017). However, understanding biochar degradation by 107 microbes is far from being complete, and the impact of biochar on soil eukaryotes remains 108 poorly known. Data on the effects of biochar on soil invertebrates are constrained to lab 109 ecotoxicological tests on selected or model animals (Marks et al., 2014; Domene et al., 110 2015) and to a few short-term field experiments with earthworms (Weyers and Spokas, 111 2011; Tammeorg et al., 2014), protists and nematodes (Eo et al., 2018) and epigeous 112 macroinvertebrates (Castracani et al., 2015) as indicators. 113 Biochar chemical and physical properties, including its resistance to microbial 114 utilization, change with aging (Mukherjee et al., 2014) which has profound implications 115 for the estimation of the long-term capacity of biochar-amended soils to sequester carbon 116 (Spokas, 2013). Biochar mineralization may be described following an exponential model 117 (Lehmann et al., 2009) with an initial phase of fast decomposition of the labile fraction 118 followed by a second phase of slow decomposition of the recalcitrant aromatic condensed 119 carbon components (Wang et al., 2016). But this model is continuously reshaped by soil- 120 biochar-soil biota interactions that also change over time (Ameloot et al., 2014). 121 However, only very recently, multiyear experiments have begun to provide data on the 122 mid-term evolution of soil microbial communities in biochar-amended agricultural soils 123 under field conditions (Nielsen et al., 2014; Watzinger et al., 2014; Mackie et al., 2015, 124 Jones et al. 2012; Yao et al., 2017; Mitchell et al., 2016). Regrettably, data about mid- 125 term effects of biochar on higher levels of the soil community are even scarcer and, to 126 our knowledge, restricted to the works of Domene et al. (2014) on soil invertebrate 127 feeding activity and Pressler et al. (2017) on several groups of the soil food web. 5 128 Based on biochar’s proven ability to alter soil chemistry and physical structure and to 129 improve soil water retention capacity, we hypothesized that biochar addition to our soils 130 will: (a) increase soil microbial biomass, (b) lead to greater abundance of soil micro- 131 arthropods and particularly of the water-dependent forms, and (c) alter the composition 132 of the soil microbial and micro-arthropods’ communities in the short- and mid-term. We 133 also posited that (d) soil microbes would feed on biochar-derived carbon at least in the 134 initial period after biochar application to soil. 135 To test these hypotheses, a maize-derived biochar was applied to the soil of a vineyard 136 and soil microbes and soil micro-arthropods were monitored during the following two 137 years in biochar-amended and control plots. Maize is a C4 plant and our experimental 138 vineyard soils historically developed under C3 vegetation. The difference between the 13C 139 isotopic signature of C4 and C3 plant-derived soil organic matter (SOM) was used to 140 monitor the inclusion of biochar-C in the diet of different soil microbial groups over time. 141 The sustainability of the biochar strategy will depend on to what extent biochar is 142 degraded and on its medium and long-term effects on the native soil biota, which are 143 ultimately responsible for soil environmental services. With this in mind, our work aimed 144 to assess the effects of biochar on the soil biota of a vineyard under semi-arid 145 Mediterranean conditions. 146 2. Material and methods 147 2.1. Work area 148 The experimental plots were located in a vineyard located in Vimbodí i Poblet 149 (Tarragona, Spain; 41º 22’ 43.8” N; 01º 04’ 30.3” E, 527 m.a.s.l.). Local topography is 150 gentle (8% slope) and soils are deep and well drained Fluventic Haploxerept (Soil Survey 151 Staff, 2014) evolved from Quaternary detrital materials. Surface and sub-surface horizons 6 152 (0-40 cm) contain great amounts of coarse elements (55% to 70%). The surface horizon 153 (0-20 cm) is sandy loam and has a neutral pH, poor cation exchange capacity (7.1 cmolc 154 kg−1) and low content of carbonates. Soil organic matter content (about 1.7%) is within 155 the normal range for agricultural soils (see other soil properties in Table 2). Climate is 156 dry continental Mediterranean, with 550 mm of total annual rainfall and 14.6°C of mean 157 annual temperature. Daily temperature and rainfall during the working period are shown 158 in Fig. S1. 159 Vines (Vitis vinifera ssp. vinifera) were planted in 1992 at a density of 4000 plants ha-1 160 with a planting pattern of 2.20 m x 1 m. The vine plants are trellised and managed with 161 double Royat pruning. Pests are controlled with copper hydroxide (50%), wettable 162 sulphur (80%), sulphur in powder (95.5%) and Spinosad (SPINTOR 480, Dow 163 Agrosciences LLC, USA), a natural insecticide obtained from Saccharopolyspora 164 spinose, commonly used in organic farming. Weeds are mechanically removed by 165 ploughing the interrow spaces of the plantation to a depth of 15 cm three to six times per 166 year. In 1990, the vineyard was fertilized with compost made of cow manure, after which 167 no other fertilizer has been applied. 168 2.2. Biochar production 169 In order to trace the fate of biochar-C in the vineyard soil, biochar from maize corn cob 170 rachis was used. Maize (Zea mais) was chosen because, being a C4 plant, its 13C isotopic 171 signature significantly differs from that of Mediterranean soils historically cultivated with 172 C3 plants (δ13C value ranges from -24 to -32 ‰ for C3 plants and from −7 to −17‰ for 173 C4 plants; Boutton,1996). This difference allows the flux of the biochar-derived carbon 174 to be followed through the belowground food web (Fry et al., 1978) by isotope analysis 175 of carbon resources and consumers. Corn cob biomass contained 30% water and was 176 pyrolyzed in the furnace of the Environmental North Valorization Center of the Touro 7 177 mine (A Coruña, Spain). The slow pyrolysis started at ambient temperature and reached 178 a final temperature of 450 to 500 ºC. The residence time of the biomass at final 179 temperature was two hours. 50 to 65% of the initial biomass-C (equivalent to 25% to 32% 180 of the initial biomass) was recovered as biochar. Biochar chemical-physical properties 181 were analyzed as described in Raya-Moreno et al., (2017) and are reported in Table 1. 182 2.3. Experimental design and sampling plan 183 In May 2013, six contiguous 90 m2 field plots (Fig. S2) were demarcated in a two- 184 hectare vineyard and set up as a field experiment following a random design with three 185 plots assigned to biochar application (Bc) and three more plots assigned to non-amended 186 controls (Co). Biochar was homogeneously spread on the soil of the Bc plots with a 187 fertilizer spreader at a dose of 5 Mg C ha-1 (equivalent to 6.5 g kg−1) and incorporated 188 into the soil by ploughing at 15 cm depth. The control plots were ploughed the same way. 189 A week after biochar application and ploughing, top soil (0-10 cm) was sampled from 190 each biochar-amended and control plot and analyzed for basic properties. All plots were 191 analyzed two, fourteen and twenty-four months after biochar application for total, 192 inorganic and organic soil carbon (Table 2). 193 From July 2013 to April 2015, we conducted two different soil sampling campaigns. To 194 evaluate the effects of biochar on soil microbial communities and on biochar-C 195 exploitation by soil microbes, the field plots were sampled seasonally for a total of eight 196 times. At each sampling date, six soil samples per plot were extracted as described below 197 and combined in pairs to produce three composite samples per plot. To measure effects 198 of biochar on soil micro-arthropod communities, we sampled the plots the first day of 199 February, May, August and November 2014 and took eight soil samples per plot each 200 time. At all times, soil samples were extracted with 5 x 5 x 15 cm soil borers from random 8 201 points situated 1 m away from each other in the central line of the four interrows between 202 vines of each plot. 203 2.4. Microbial phospholipid fatty acid (PLFA) extraction and isotopic ratio determination 204 PLFAs were microwave-extracted from freeze-dried soils with a 0.1 M phosphate 205 buffer:choloform:methanol solution at a 0.8:1:2 ratio. For the quantification and 206 identification of PLFAs, 20 μl of 19:0 phosphatidylcholine (Avanti Polar Lipids Inc., 207 Alabaster, AL) were added as internal standard. Lipids were extracted and partitioned 208 into glycolipids, phospholipid and neutral lipids and phospholipids were transesterificated 209 to obtain fatty acid methyl esters (FAMEs). FAMEs were analyzed with capillary gas 210 chromatography with flame ionization detector (GC-FID 7820A, Agilent Technologies, 211 Palo Alto, USA) with a HP1-MS capillary column (60 m x 0.25 mm x 0.25 μm film 212 thickness). The program started at 80ºC, followed by a heating rate of 10ºC minute-1 to 213 170 ºC, 2 ºC minute-1 to 230 ºC, 5 ºC minute-1 to 310 ºC, with a final hold of 10 minutes. 214 FAMEs were identified and quantified from mass spectral and retention time matches to 215 the NIST 2008 mass spectral library. The isolated PLFAs were grouped into biomarkers 216 of microbial groups as shown in Table 3. 217 Effects of biochar on the diet of soil microbial groups were explored by comparing the 218 δ13C signature of their specific PLFA biomarkers in the biochar-amended and control 219 soils. The δ13C unit was used to report 13C isotope data as in Craig (1953): 𝑅𝑠𝑎𝑚𝑝𝑙𝑒 − 𝑅𝑃𝐷𝐵 220 δ13 C = 1000 ‰ 𝑅𝑃𝐷𝐵 221 The δ13C signature and carbon content of the most significant FAMEs were analyzed 222 (only microbial markers present in samples in sufficient concentration over time were 223 taken into account) by capillary gas chromatography-combustion-isotope ratio mass 224 spectrometry (GC-C-IRMS) (Trace GC Ultra, GC-C Combustion III and DeltaV IRMS, 9 225 Thermo Scientific, Bremen, Germany). FAME separation was performed with a capillary 226 GC column type DB-5 (length 60 m, i.d. 0.25 mm, film thickness 0.25 μm; Agilent 227 Technologies, Santa Clara, CA). The GC temperature programme started at 80 ºC with a 228 1-minute pause, followed by a heating rate of 10 ºC minute−1 to 170 ºC, 2 ºC minute−1 to 229 230 ºC and 5 ºC minute−1 to 310 ºC, with a final pause of 10 minutes. The δ13C values 230 were corrected by using working standards (18:0 and 24:0) calibrated on an elemental 231 analyser-IRMS (Flash 1112, Thermo Scientific, Bremen, Germany) coupled to a DeltaV 232 IRMS continuous flow IRMS (Thermo Scientific, Bremen, Germany). The final δ13C 233 values of the PLFAs, were obtained after correcting the measured δ13C FAME values for 234 the addition of the methyl group during transesterification by simple mass balance (after 235 Denef et al., 2007): 𝑁 𝑥 δ13 CPLFA−Me − δ13 CMeOH 236 δ13 CPLFA = (𝑁 − 1) 237 where the isotope value of the PLFA (δ13CPLFA) was calculated from the isotope ratio of 238 the PLFA methyl ester (δ13CPLFA-Me), the isotope ratio of methanol used for methylation 239 (δ13CMe-OH), and the number of C atoms of the methylated PLFA (N). 240 To evaluate a possible effect of biochar on the efficiency of PLFA extraction from soil 241 samples (Gomez et al., 2014), C19:0 PLFA was added in a known dose to three soil 242 samples taken from the non-amended control plots and to three more samples taken from 243 the biochar-amended plots. All samples were processed for PLFA extraction as explained 244 above. Results were corrected for extraction efficiency (EE), calculated as the percentage 245 of C19:0 recovered relative to the dose added. EE was 72.7% in the non-amended control 246 soils and 71.7% in the biochar-amended soils. 247 2.5. Determination of soil micro-arthropod community size and composition 10 248 Soil micro-arthropods were heat extracted from soil samples with 70% ethyl alcohol 249 using Tullgren funnels (Moore et al., 2000) during eight days. The extractors were 250 operated in the dark during the first two days to prevent mortality due to fast soil drying. 251 The animals collected where classified under the microscope to different taxonomic 252 levels, counted and classified into functional groups based on common food preferences 253 and life traits (Table S1). 254 2.6. Data analysis 255 The sum of all biomarkers (in nMol PLFA g−1 soil) was used as a proxy for total 256 microbial biomass. The fungal to bacterial biomass ratio was calculated by dividing the 257 biomass of the fungal PLFA 18:2ω6,9c by the sum of PLFAs a15:0, i16:0, i17:0, a17:0, 258 16:1ω7t, 16:1ω7c, 17:1ω7c, 17:0cy, 19:0cy, 18:1ω5c, 14:0, 15:0, 17:0 and 18:0. 259 Effects of biochar amendment and time after soil amendment on microbial biomass, 260 abundance of micro-arthropods, fungal-to-bacterial biomass ratio and PLFA isotopic 261 signature were tested according to a mixed model design, with “treatment”, “time” and 262 their interaction as fixed factors and “plot” as a random factor. The analyses were 263 performed with the lmer function of the lme4 package (Bates et al., 2015) in R (R 264 Development Core Team, 2016). Tests for fixed effects were done with the lmerTest 265 package (Kuznetsova et al., 2016) with the Kenward-Roger's approximation for 266 denominator degrees of freedom for F (Kenward and Roger, 1997). Tests for differences 267 between treatment levels after fitting the linear models were evaluated from predicted 268 marginal means using the lsmeans package (Lenth, 2016). 269 Effects of treatment and time on the communities of soil microbes and micro-arthropods 270 were studied by permutational analyses of variance (PERMANOVA) and were 271 graphically represented using distance-based redundancy analyses (dbRDA). The 272 contribution of each group of microbes or micro-arthropods to dissimilarity between 11 273 samples was evaluated by SIMPER analyses. PERMANOVAs and dbRDAs were 274 performed with PERMANOVA+ for PRIMER (Anderson et al., 2008), and SIMPER 275 analyses with PRIMER v.7 (Clarke and Gorley, 2015). 276 3. Results 277 3.1 Physical and chemical characteristics of the experimental soils 278 Biochar addition doubled the carbon content of the experimental soils (from 10.7 g C 279 kg soil-1 in the control plots to 21.3 g C kg soil-1 in the biochar amended plots). This 280 increase particularly affected the non-soluble and non-oxidizable fractions of the soil 281 carbon pool that amounted to 9.4% of total carbon in the control soils and to 26.7% in the 282 amended soils (Table 2). The soil pH increased from 7.3 to 7.7 and the C/N ratio from 283 15.3 to 26.6. Electrical conductivity almost doubled after biochar addition although the 284 new value (0.14 dS m-1) remained below the adequate salinity threshold for agricultural 285 soils. There was no change in total soil carbon in the control plots over time. In contrast, 286 during the two-year monitoring period, there was a 22% reduction of the total soil carbon 287 content in the biochar amended soils, due to the decline of both the organic and the 288 inorganic fractions of the carbon pool (22.6% and 14.3% respectively) (Table 2). 289 3.2. Effects of biochar on soil microbial biomass 290 Biochar had either no effect or a negative effect on total soil microbial biomass 291 depending on time (Treatment x Time: p = 0.0005). Microbial PLFA content was 292 significantly lower in the biochar-amended soils than in the control soils on most 293 sampling dates (Fig. 1a). The greatest difference between control and biochar-amended 294 soils (85.3 ± 1.2 and 32.2 ± 1.2 nMol PLFAs g-1 soil respectively) occurred in April 2014. 295 Only on two dates (November 2013 and April 2015), when microbial biomass was very 12 296 low in the control soils (2.6 ± 1.2 and 7.5 ± 1.2 nMol PLFAs g-1 soil respectively), was 297 the effect of biochar insignificant. 298 3.3. Effects of biochar on the structure of the microbial community 299 The effect of biochar on the fungal-to-bacterial biomass ratio depended on time 300 (Treatment x Time: p = 0.037). The ratio was affected by the treatment only in January 301 2014 and was significantly lower in the biochar-amended soils (0.04 ± 0.01) than in 302 controls (0.11 ± 0.01) (Fig. 1b). 303 The composition of the soil microbial community, as indicated by the proportion of 304 individual PLFA microbial biomarkers, significantly depended on time (PERMANOVA: 305 p = 0.0001) but was not affected by the addition of biochar (Table S2 and Fig. 2a). In the 306 dbRDA, 63.1% of the variance was explained by Axis I along which the samples were 307 ranked by sampling dates. Sampling T4 and T5, located towards the right side of the axis, 308 were done immediately after rainy periods while sampling T2, in the opposite end of the 309 axis, was done after an extended dry period (Fig. S1). A regression analysis showed that 310 the scores of the samples on Axis 1 were positive and significantly related with total 311 microbial biomass (R² = 0.4834; p < 0.0001) and with the fungi-to-bacteria ratio (R² = 312 0.505; p < 0.0001). The SIMPER analysis showed that the main contributors to the 313 formation of Axis I were the universal microbial marker 16:0, the fungal marker 314 18:2ω6,9c and the gram-negative marker 16:1ω7c. 315 3.4. Soil micro-arthropod community abundance and composition 316 Biochar did not alter the total abundance of soil micro-arthropods. Their abundance only 317 depended on time (p = 0.0007): they were significantly (p < 0.05) more abundant (24,121 318 individuals m-2) in the spring sampling (May 2014) than in any other sampling date 319 (10,686 in winter -February 2014-; 13,467 in summer -August 2014-; 12,915 in fall – 320 November 2014). 13 321 Biochar had no effect on the composition of the micro-arthropod community (Table S2). 322 The community composition only significantly changed over time (PERMANOVA: p = 323 0.0001). The dbRDA graph showed the samples grouped by sampling date along Axis 1, 324 with the spring and winter samples located in opposite sides along the axis (Fig. 2b). A 325 SIMPER analysis showed that differences between samples were mainly due to 326 endeostigmatic mites and immature oribatids that were more abundant in spring and 327 summer than in winter or fall. 328 3.5. Isotopic signature of the PLFA microbial markers 329 δ13C was -13.12 ± 0.01 for the maize biochar and -26.84 ± 0.05 for soil. The signature 330 of the non-amended soil was measured three times (in 2013, 2014 and 2015) and changes 331 over time were not significant. 332 Twelve microbial PLFAs were extracted from the soil samples in sufficient quantity to 333 allow the measurement of their isotopic signature (Table S3) although some of them were 334 not present in all samples. There were significant (p < 0.05) differences in mean annual 335 isotopic values between PLFA types: 18:1ω5c (δ 13C = -19.9 ± 1.7) was the most 13 C 336 enriched PLFA, followed by PLFAs a15:0, 15:0 and 16:1ω7c (δ 13C = -24.9 ± 0.3) and 337 by the remainder PLFAs, with δ 13C values between -27.1 and -31.2. 338 The signature of all PLFAs varied significantly over time (Fig. 3). Moreover, biochar 339 modified significantly the isotopic signature of four PLFAs (15:0, 16:1ω7c, 16:0, and 340 18:0). In the four cases PLFAs were enriched in 13C in the biochar-amended soils relative 341 to controls (Fig. 4 and Table 4). Two more PLFAs (a15:0 and i16:0) were sensitive to the 342 interaction between time and treatment in such a way that they were 13C enriched by the 343 addition of biochar compared to control only in sampling T2 (Fig. 5 and Table 4). 344 4. Discussion 345 4.1. Effects of biochar on soil microbial biomass 14 346 We had hypothesized that amendment with biochar will increase soil microbial 347 biomass, but this did not happen. In the control plots, microbial biomass evolved 348 according to the phenology of the vine plants, with a spring peak from March to 349 September when vines are active and labile carbon is provided to soil microbes by roots 350 (Amendola et al., 2017), low values from November to March, during the plant dormancy 351 period in the region (Camps et al., 2012) and minima occurring during drought. Biochar 352 had no effect on the winter basal soil microbial biomass but suppressed its spring peak. 353 Given the great sorption capacity of biochar (Wang et al., 2010), the suppressive effect 354 could be attributable to a reduction of available resources due to sorption of the spring 355 labile rhizodeposition carbon onto biochar surfaces (Foster et al., 2016). Positive effects 356 on soil microbial biomass have been previously reported after alkaline biochar application 357 to acidic soils (Pragoyo et al., 2010; Paz-Ferreiro et al., 2011; Ameloot et al., 2013; 358 Mackie et al., 2015; Jiang et al., 2016; Zheng et al., 2016). However, our results are in 359 line with those of a number of studies that show no effect (Castaldi et al., 2011) or reduced 360 microbial biomass when biochar is applied to neutral or alkaline soils (Warnock et al., 361 2010; Luo et al., 2011; Dempster et al., 2012; Lu et al., 2014; Ameloot et al., 2014). 362 4.2. Effects of biochar on the soil microbial community 363 We expected that biochar application would cause changes in the relative abundance of 364 soil microbial functional groups but, surprisingly, neither the fungi-to-bacteria ratio nor 365 the overall microbial community composition were significantly affected. A number of 366 previous studies have shown that biochar selectively stimulates microbes involved in 367 nitrogen cycling and phosphate solubilization (Ducey et al., 2013), gram-positive and 368 gram-negative bacteria (Gomez et al., 2014), actinomycetes (Prayogo et al., 2014) and 369 fungi (Steinbeiss et al., 2009). These effects are often explained by changes in quantity 370 and quality of available nutriments: the labile biochar fraction might benefit copiotrophic 15 371 against oligotrophic bacteria (Xu et al., 2016; Jenkins et al., 2017) while the recalcitrant 372 fraction might favor oligotrophic and gram-positive bacteria (Ding et al., 2013). The lack 373 of response of our microbial groups to biochar application may be due to the lower 374 agronomic dose at which biochar was applied compared to most experiments (Ameloot 375 et al., 2013, Abujabhah et al., 2018, Watzinger et al., 2014). In the same sense, this 376 moderate dosage resulted in minor changes in soil pH, which is the main driver of changes 377 in soil microbial community composition (Fernández-Calviño et al., 2010; Rousk et al., 378 2010; Anders et al., 2013; Farrell et al., 2013). 379 4.3. Effect of biochar on soil micro-arthropods 380 We had also postulated that biochar would increase the abundance of micro-arthropods 381 at different trophic levels of the soil food web, based on an expected increase of basal 382 resources and microbial preys (Abujabhah et al., 2016). But instead, the abundance of the 383 targeted fauna was not affected by biochar. Our results are in line with those of Domene 384 et al. (2014) and Pressler et al. (2017) who did not find effects of biochar on the feeding 385 activity of the soil fauna nor on the abundance of any functional group of the soil food 386 web, respectively. Despite the paucity of data for comparison, it seems that the response 387 of soil micro-arthropods to biochar application follows the same trend as soil microbial 388 biomass: a positive response in acidic soils (as in Abujabhah et al., 2016) and no response 389 in neutral to alkaline and hard-textured soils (as in Domene et al., 2014, Pressler et al., 390 2017). In the same sense, we had anticipated biochar addition to favor soil water- 391 demanding invertebrates (in particular collembolans) given the biochar ability to improve 392 soil water retention under arid and semi-arid conditions (Novak et al., 2012; Obia et al., 393 2016), but this effect was not observed. Again, this could be due to the moderate rate at 394 which biochar was applied. 395 4.4. Microbial utilization of biochar and other soil resources 16 396 We assumed that SOM and biochar were the only two providers of carbon to the soil 397 food web and that the isotopic signature of the microbial PLFAs would lay between the 398 δ13C values of soil (-26.84 ‰) and biochar (-13.12 ‰). Unexpectedly, most PLFAs 399 showed δ13C values far below those of the bulk SOM. Such negative values may be due 400 to two main causes: 13C fractionation by microbes and selective preferences of microbial 401 groups towards diverse carbon sources. There is growing evidence that microbial isotopic 402 fractionation may be significant (Henn et al., 2002) and that relative deviation of δ13C 403 values of individual PLFAs compared to the δ13C value of bulk SOM may be remarkable 404 (Glasser, 2005). Several experiments have shown that microbial PLFAs can be strongly 405 depleted in δ13C (to up to 17 ‰) compared to the exploited substrate depending on 406 microbial metabolism and environmental conditions (Abraham et al., 1998; Burke et al., 407 2003; Ruess et al., 2005). For example, isotopic fractionation between bulk SOM and 408 PLFA 16:0 has been shown under anaerobic conditions (Cifuentes and Salata, 2001). Our 409 very negative PLFA δ13C values might reveal the existence of carbon sources other than 410 SOM and biochar (Williams et al., 2006) as well as microbial preferences for specific 411 fractions of the SOM (Schweizer et al., 1999; Ehleringer et al., 2000) and in particular for 412 those rich in lignin and lipids that are 13C depleted relative to sugar, starch and cellulose 413 (Bowling et al., 2008). Comparable results have been provided by Kramer and Gleixner 13 414 (2008) who found soil fungal PLFA biomarkers C depleted compared to bulk SOM, 415 probably due to preferential exploitation of lignin (Glaser, 2005). 416 In our vineyard, the diet of all soil microbial groups changed over time, most likely 417 following shifts in resource availability. In this sense, the lowest δ13C values were found 418 for all groups in the 2015 winter sampling that was carried out a few days after the pruning 13 419 of the vine trees, when lignified (and therefore C depleted) vine cuttings fell to the 17 420 ground. The only exception to this trend was provided by the 18:1ω5c biomarker that 421 indicates a gram-negative group that clearly prefers more 13C enriched resources. 422 We found evidence that biochar alters the diet of several soil bacteria. In four cases (a 13 423 gram-negative marker and three general bacterial markers), the biomarkers were C 424 enriched in the biochar amended soils relative to controls throughout the whole 13 425 monitoring period. Although being significant, the C enrichment was constant and 426 always low (Δ ≈ 1‰ -2‰) which indicates little importance of biochar in the diet of the 427 soil bacteria (and negligible use by fungi) during the two years following soil amendment. 428 It has been reported that the incorporation of biochar-derived carbon in microbial 429 biomass starts immediately after biochar application to soil. The initial and very active 430 decomposition phase last from two days to two months (Kuzyakov et al., 2009; Smith et 431 al., 2010; Santos et al., 2012), after which the mineralization rate stabilizes at low levels 432 and microbes continue to consume small doses of biochar carbon for many years 433 (Kuzyakov et al., 2014). Since in this work the first sampling was carried out two months 434 after biochar application, a potential short period of fast mineralization might have been 435 missed, but another explanation for the low incorporation of biochar in microbial biomass 436 might be the quality of the biochar. Some authors have found maize biochar more 437 recalcitrant than biochar made of other feedstocks due to the presence of strong structural 438 surface functional groups (Purakayastha et al., 2015). Biochar lability is directly related 439 to oxygen content, and the very low O/C ratio of our biochar (0.1) might have contributed 440 to its recalcitrance, since biochars with O/C < 0.2 have half-lives in the range of thousands 441 of years (Spokas et al., 2010). 442 A very interesting finding is that the 13C signature of gram-positive bacteria markers 443 suggested that this group of bacteria only included biochar in its diet in November 2013, 444 when microbial biomass was at its lowest after an extremely dry summer. For two 18 13 445 biomarkers (a15:0 and i16:0), the measured C enrichment (2.5‰ and 3.4‰ 446 respectively), i.e. the proportion of biochar-derived carbon in their diet, was higher than 447 the observed for any other bacterial group at any other sampling date. This suggests that 448 biochar-C might be a food resource when no other option is available, at least for 449 microbes able to decompose recalcitrant aromatic soil carbon as is the case of gram- 450 positive bacteria (Kramer and Gleixner, 2008). This is in accordance with Jiang et al. 451 (2016) who found that soil microorganisms prefer to use SOC sources more labile than 452 biochar, when available. However, this interpretation must be used with caution, because 13 453 summer drought stimulates the production of C depleted biomass by plant roots 454 (Bowling et al., 2008) and the observed shift in the two bacterial PLFAs might be the 455 consequence of microbial consumption of rhizodeposition products more than of biochar. 456 5. Conclusion 457 This study demonstrated that corn cob biochar applied at agronomic doses to 458 Mediterranean neutral to alkaline soils can reduce soil microbial biomass for at least two 459 years after application. At this application rate, biochar had no significant effects on the 460 community composition of soil microbes or micro-arthropods. Microbial utilization of 461 biochar was very low which is promising in order to increase the residence time of 462 biochar-derived carbon in soil. The isotopic signature of PLFA biomarkers indicated that 463 in our soil, microbes feed preferably on organic matter fractions more 13C depleted than 464 the supplied biochar. However, under the severe conditions of Mediterranean summers, 465 biochar might constitute an emergency resource for soil microbes to overcome food 466 shortage during drought. 467 Acknowledgements 19 468 This work was funded by the MEDICHAR project (ref. AGL2012-40037-C02-01) of 469 the Spanish Ministry of Economy and Competitiveness. A.R.-M- acknowledges support 470 from the MINECO Maria de Maeztu award MDM-2015-0552. The authors wish to 471 express their most sincere gratitude to Dr. Felipe Macías Vazquez (Universidad de 472 Santiago de Compostela) for his invaluable collaboration in growing the maize and 473 preparing the maize biochar and Dr. P. Comes and N. Moraleda for laboratory assistance. 474 We are also very grateful to Miguel Torres Winery that provided the experimental 475 vineyard and precious information about soil characteristics and vineyard management. 476 References 477 Abraham, W.-R., Hesse, C., Pelz, O., 1998. Ratios of carbon isotopes in microbial lipids 478 as an indicator of substrate usage. Appl. Environ. Microbiol. 64, 4202-4209 479 Abujabhah, I. S., Bound, S. A., Doyle, R., Bowman, J. P., 2016. Effects of biochar and 480 compost amendments on soil physico-chemical properties and the total community 481 within a temperate agricultural soil. Appl. Soil Ecol. 98, 243-253. 482 https://doi.org/10.1016/j.apsoil.2015.10.021 483 Abujabhah, I. S., Doyle, R. B., Bound, S. 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Asterisks indicate 4 significant differences between control and biochar-amended soils at each sampling date 5 after ANOVA on log transformed data. ***: p = 0.001; **: p = 0.01; *: p = 0.05. Vertical 6 bars denote standard errors of the mean (n=3). 7 Figure 2. dbRDAs on effect of biochar addition on microbial PLFA biomarkers (a) and 8 on micro-arthropod groups (b) over time. In (a) T1: July 2013, T2: November 2013, T3: 9 January 2014, T4: April 2014, T5: July 2014, T6: October 2014, T7: January 2015, T8: 10 April 2015. In (b), T1: January 2014, T2: May 2014, T3: August 2014, and T4: November 11 2014. 12 Figure 3. Mean δ13C value of the soil microbial PLFA biomarkers over time. For each 13 sampling date, the mean was calculated from all samples (including all biochar-amended 14 and non-amended plots). Vertical bars denote standard errors of the mean (n=3). 15 Figure 4. Effect of biochar on the isotopic signature of four microbial PLFAs. 16 Significance of differences between control soils (Co) and biochar-amended soils (Bc) 17 after ANOVA on transformed data (after Ln -δ13C). ***: p = 0.001; **: p = 0.01; *: p = 18 0.05. Vertical bars denote standard errors of the mean (n=3). 19 Figure 5. Isotopic signature of PLFAs a15:0 (a) and i16:0 (b) in control soils (Co) and in 20 biochar amended soils (Bc) over time (i16:0 concentration in samplings T3, T4 and T5 21 was too low for isotopic analysis). Significance of the difference between control and 22 biochar-amended soils after ANOVA on transformed data (after Ln -δ13C). ***: p = 23 0.001; **: p = 0.01; ns = no significant difference. Vertical bars denote standard errors of 24 the mean (n=3). In (a) T1: July 2013, T2: November 2013, T3: January 2014, T4: April 1 25 2014, T5: July 2014, T6: October 2014, T7: January 2015, T8: April 2015. In (b), T1: 26 January 2014, T2: May 2014, T3: August 2014, and T4: November 2014. 2 Fig. 1 Fig. 2 20 (a) C C C C B BBC C C C 16:1ω7c 18:2ω6,9c Axis II; 14.6% 16:1ω7t C C BC C C B BB 0 B BC B B B B C BCC Time C B B CB C B B T1 BB C T2 B C C 16:0 B T3 T4 18:0 T5 C T6 T7 B T8 -20 -20 0 20 40 50 Axis I: 63.1% 40 (b) Poduromorpha C B C C B C 20 B B C Immature Cryptostigmata Axis II: 17.5% C B C 0 B Endeostigmata B C B B C C B B C -20 C Time B T1 T2 T3 T4 -40 -40 -20 0 20 40 60 Axis I: 36.4% Fig. 3 Fig. 4 C16:0 C15:0 C16:1ω7c C18:0 -10 -15 -20 δ13C -25 *** -30 * ** ** Bc Co -35 Fig. 5 Bc Co a) T1 T2 T3 T4 T5 T6 T7 T8 0 -5 -10 δ13C -15 -20 -25 ns ns ns ns ns ns ** ns -30 b) T1 T2 T6 T7 T8 0 -5 -10 δ13C -15 -20 -25 ns ns ns -30 *** ns Table 1. Elemental analysis, molar ratios and chemical properties of biochar. LOI: weight loss- on-ignition, LPO: weight loss-on-peroxide oxidation, sO: organic carbon destroyed by strong potassium dichromate oxidation, mO: organic carbon destroyed by mild potassium dichromate oxidation, AH: organic carbon resistant to acid hydrolysis. BIOCHAR PROPERTIES pH (H2O, 1:20) † 10.3 ± 0.04 δ13C ‰ -13.12 EC dS m-1 (1:5, 25 ºC) † 2.5 ± 0.5 Exchangeable Ca (meq 100 g-1) 2.02 Total C (g kg-1) † 785.8 Exchangeable Mg (meq 100 g-1) 1.1 Inorganic C (g kg-1) † 2.7 ± 0.1 Exchangeable K (meq 100 g-1) 85.3 Total N (g kg-1) † 6.8 Exchangeable Na (meq 100 g-1) 0.46 Total H (g kg-1) † 19.1 Tot P (mg kg-1) 1,838 Total S (g kg-1) † 0.64 Total Al (mg kg-1) 1,020 O (g kg-1) † 89.4 Total Fe (mg kg-1) 7,810 Ash (g kg-1) † 91.1 Total Na (mg kg-1) 200 H/C † 0.29 Total K (mg kg-1) 23,400 O/C † 0.11 Total Ca (mg kg-1) 2,550 LOI (g kg-1) † Total Mg (mg kg-1) 1,100 375 ºC 891.7 ± 0.3 Total Cu (mg kg-1) 52 550 ºC 897.9 ± 0.2 Total Co (mg kg-1) 10 950 ºC 917.7 ± 0.2 Total Cr (mg kg-1) 17 LPO (g kg ) † -1 19.5 ± 3.8 Total Ni (mg kg-1) 38 sO (g kg-1) † 235.3 ± 40.1 Total Pb (mg kg-1) 13 mO (g kg-1) † 43.7 ± 3.6 Total V (mg kg-1) 10 AH (g kg-1) † 65.7 ± 8.5 Total Zn (mg kg-1) 410 Particle sizes (% d. w.) Total As (µg kg-1) 396 5-2 mm 2.8 Total Cd (µg kg-1) 38 2-1 mm 40.1 PAHs (16 US EPA, mg kg-1) 40 1-0.5 mm 24.6 0.5-0.2 mm 27.3 <0.2 mm 5.2 † Data from Raya-Moreno et al., 2017. Table 2. Selected soil characteristics in control plots and in plots amended with biochar one week after biochar application. Data correspond to the top 10 cm of the soil and are reported as mass ratio in the < 2mm soil fraction (except stoniness). Mean ± Stdev (n=3). Biochar-amended Control plots plots One week after biochar application Stoniness (% of field sample) 61.7 ± 4.2 64.4 ± 3.9 Sand (%) 57.7 ± 2.6 60.6 ± 1.3 Loam (%) 26.9 ± 1.9 23.7 ± 2.0 Clay (%) 14.8 ± 0.6 15.3 ± 0.5 δ13C ‰ -26.84 -19.81 pH (water 1:2.5) 7.3 ± 0.4 7.7 ± 0.3 EC μS cm-1 (1:5, 25ºC) 76.3 ± 15.4 140.1 ± 29.2 Total C (g kg-1) 10.7 ± 0.8 21.3 ± 1.5 Oxidizable C (g kg-1) 9.6 ± 0.6 15.5 ± 0.4 -1 Soluble C (1:2.5 mg kg ) 90.7 ± 33.3 115.4 ± 24.7 N (Kjeldahl, %) 0.07 ± 0.001 0.08 ± 0.006 P (Olsen, mg kg-1) 12.7 ± 2.1 18.3 ± 1.5 CaCO3 (g kg-1) 0.95 ± 0.2 1.12 ± 0.9 -1 CEC (meq 100 g ) 7.1 ± 1.2 7.3 ± 0.4 Exchangeable Ca (meq 100 g -1) 6.1 ± 1.2 5.8 ± 0.5 -1 Exchangeable Mg (meq 100 g ) 0.5 ± 0.1 0.5 ± 0.1 -1 Exchangeable K (meq 100 g ) 0.4 ± 0.02 0.9 ± 0.1 Exchangeable Na (meq 100 g -1) < 0.07 < 0.07 Two months after biochar application Total C (g kg-1) † 10.72 ± 0.79 21.33 ± 1.50 Total inorganic C (g kg-1) † 0.94 ± 0.25 1.12 ± 0.95 Total organic C (g kg ) † -1 9.77 ± 0.54 20.21 ± 2.37 14 months after biochar application Total C (g kg-1) † 11.41 ± 0.86 18.53 ± 2.98 Total inorganic C (g kg-1) † 1.43 ± 0.99 1.35 ± 0.37 Total organic C (g kg ) † -1 9.99 ± 1.03 17.15 ± 3.31 24 months after biochar application Total C (g kg-1) † 10.29 ± 0.73 16.60 ± 1.03 Total inorganic C (g kg ) † -1 0.49 ± 0.29 0.96 ± 1.11 Total organic C (g kg-1) † 9.80 ± 0.85 15.64 ± 2.03 † Data from Raya-Moreno et al., 2017. Table 3. Microbial functional groups assigned to phospholipid fatty acid (PLFAs) biomarkers. Functional group PLFA markers References Gram-positive bacteria a15:0, i16:0, i17:0, a17:0 Frostegård & Bååth (1996), Zelles (1997) Gram-negative bacteria 16:1ω7t, 16:1ω7c, 17:1ω7c, 17:0cy, 19:0cy, 18:1ω5c Frostegård & Bååth (1996), Zelles (1997) Actinomycetes 10Me16:0, 10Me18:0 Ringelberg et al. (1997) Saprophytic fungi 18:2ω6,9c Frostegård & Bååth (1996), Bossio & Scow (1998) Non-specific bacterial 14:0, 15:0, 17:0, 18:0 Bossio & Scow (1998) Universal microbial 16:0 Bossio & Scow (1998) Table 4. Two-way ANOVA (GLM procedure) for the 13C isotopic signatures of all microbial PLFA markers depending on treatment (soil vs soil + biochar) and time (eight sampling dates over two years). Only significant interactions are shown. DF F value Pr (>F) 18:1ω5c Treatment 1 13.049 0.2668 Time 4 71.411 0.0009 *** a15:0 Treatment 1 130.482 0.0012 ** Time 7 81.336 0.0238 * Treat x Time 7 26.553 0.0313 * 15:00 Treatment 1 14.376 0.0005 *** Time 7 9.508 0.0012 ** 16:1ω7c Treatment 1 71.934 0.0109 * Time 7 56.563 0.0001 *** 16:1ω7t Treatment 1 14.368 0.2383 Time 7 97.889 0.0007 *** i16:0 Treatment 1 32.637 0.0320 * Time 4 21.483 0.0027 ** Treat x Time 4 5.948 0.0039 ** 10Me-16:0 Treatment 1 28.796 0.0983 Time 7 42.94 0.0015 ** 10Me-18:0 Treatment 1 0.2915 0.5953 Time 4 101.713 0.0001 *** 16:00 Treatment 1 121.828 0.0012 ** Time 7 42.928 0.0014 ** 18:00 Treatment 1 76.842 0.0087 ** Time 7 71.71 0.0221 * 18:2ω6,9c Treatment 1 18.901 0.1790 Time 6 36.78 0.0070 ** 19:0cy Treatment 1 0.308 0.5824 Time 7 75.498 0.0133 * Figure S1. Mean daily temperature (MT) and total daily precipitation (PPT24h) during the study period. Data provided by the Montblanc automatic weather station (41º 22' 25'' N; 1º 9' 48'' E). Figure 2. General view of the experimental field plots. Table S1. Micro-arthropods found in the soil of the experimental plots and their main food preferences after Muraoka & Ishibashi (1976). Walter et al. (1986). Walter (1988). Behan-Pelletier (1999). Gerson et al. (2003). Krantz & Walter (2009). Walter & Proctor (2013). Castilho et al. (2015) and Van Leeuwen (2016). ACARI Endeostigmata Nanorchestidae spp. Predators Alycidae Fungivores Sphaerolichida Sphaerolichidae Predators Oribatida Phthiracarus sp.1 Polyphages Brachychthonius sp.1 Polyphages Liochthonius sp.1 Polyphages Cosmochthonius sp.1 Polyphages Eohypochthonius sp.1 Polyphages Epilohmannia sp.1 Polyphages Papillacarus sp.1 Polyphages Nothrus sp.1 Polyphages Microzetes sp.1 Polyphages Oppiidae spp Polyphages Suctobelbidae sp.1 Polyphages Tectocepheus velatus Polyphages Oribatula tibialis Polyphages Ceratozetes sp.1 Polyphages Liebstadia sp.1 Polyphages Sheloribatidae sp.1 Polyphages Sheloribatidae sp.2 Polyphages Immature Oribatida Polyphages Astigmata Acaridae Fungivores/Nematophages Hipopus forms Inactive Prostigmata Eupodidae Fungivores Anystidae Predators Scutacaridae Fungivores Tydeidae Fungivores/ Predators/Microphytophages Paratydeidae Predators Tarsonemidae Fungivores Rhagidiidae Predators on arthropods Penthalodidae Phytophages Raphignathidae Predators Stigmaeidae Predators on arthropods Cunaxidae Predators on nematodes Erythraeidae Predators Trombididae Predators Mesostigmata Ascidae sp.1 Predators Rhodacaridae spp. Predators Parasitidae spp. Predators Veigaiidae Predators on arthropods Uropodidae Fungivores Zerconidae Fungivores Immature Mesostigmata Predators MYRIAPODA Chilopoda (Geophilomorpha) Predators Symphyla Root-feeders/saprophages Pauropoda Fungivores INSECTA Protura Fungivores Diplura Diplura (Japygidae) Polyphages (mainly predators) Diplura (Campodeidae) Polyphages Collembola Poduromorpha Fungivores/Nematophages Entomobryomorpha Fungivores Symphypleona Fungivores Psocoptera Polyphages REFERENCES Behan-Pelletier, V. 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The soil life cycle: food webs and ecosystem services during soil transformations. Ph. D. thesis, Wageningen University, Wageningen, NL. ISBN 978-94-6257-626-1. Walter, D. E., 1988. Nematophagy by soil arthropods from the shortgrass steppe, Chihuahuan desert and Rocky Mountains of the Central United States. Agric. Ecosyst. Environ. 24, 307-316. https://doi.org/10.1016/0167-8809(88)90074-6 Walter, D. E., Proctor, H., 2013. Mites: Ecology, Evolution & Behaviour. CAB International. Wallingford, UK. ISBN 0-88199-375-3 Walter, D. E., Hudgens, R. A., Freckman, D. W., 1986. Consumption of nematodes by fungivorous mites, Tyrophagus spp. (Acarina: Astigmata: Acaridae). Oecologia 70, 357-361. https://doi.org/10.1007/BF00379497 Table S2. Effect of biochar and time after biochar application to soil on soil microbial and soil micro-arthropod communities after PERMANOVA. Tr: treatment; Ti: time, Pl: plot. (***, significant at P > 0.001) Unique Source df SS MS Pseudo-F P(perm) perms Microbial community Tr 1 570.32 570.32 2.3866 0.1975 10 Ti 7 8004.5 1143.5 8.5555 0.0001*** 9924 Pl (Tr) 4 955.86 238.97 1.7879 0.0624 9929 Tr x Ti 7 890.98 127.28 0.95231 0.5235 9907 Res 28 3742.4 133.66 Total 47 14164 Micro-arthropod community Tr 1 897.59 897.59 1.0671 0.3985 10 Ti 3 16,875 5,625 5.1452 0.0001*** 9923 Pl(Tr) 4 3364.7 841.16 0.76942 0.7783 9906 Tr x Ti 3 2317.4 772.45 0.70657 0.79 9913 Res 12 13119 1093.2 Total 23 36574 Table S3. Isotopic signatures (δ13C) of the PLFA microbial markers depending on treatment (Co: control soils; Bc: biochar amended soils) and sampling dates (T1 to T8 in Fig. 1). PLFAs in (A) and (B) were affected by time or by biochar amendment (treatment) of by both (treatment and time) independently; PLFAs in (C) were affected by the interaction “Treatment x Time”. Mean ± Std. Error (n=3). (A) 16:1ω7t 19:10cy 10Me-16:0 18:1ω5c 18:2ω6.9c 10Me-18:0 15:0 16:0 16:1ω7c 18:0 T1 -27.69 ± 1.02 -31.00 ± 1.02 -26.88 ± 1.02 -21.51 ± 1.04 -29.38 ± 1.04 -27.19 ± 1.02 -25.00 ± 1.01 -29.03 ± 1.01 -26.02 ± 1.01 -28.61 ± 1.02 T2 -28.49 ± 1.02 -30.18 ± 1.02 -26.69 ± 1.02 -20.78 ± 1.04 -27.54 ± 1.02 -25.22 ± 1.01 -29.37 ± 1.01 -25.08 ± 1.01 -28.96 ± 1.02 T3 -26.97 ± 1.02 -28.49 ± 1.02 -26.35 ± 1.02 - -27.25 ± 1.04 - -24.80 ± 1.01 -29.01 ± 1.01 -24.34 ± 1.01 -27.96 ± 1.02 T4 -28.22 ± 1.02 -30.89 ± 1.02 -25.96 ± 1.02 - -28.07 ± 1.04 - -24.83 ± 1.01 -29.50 ± 1.01 -24.23 ± 1.01 -29.85 ± 1.02 T5 -28.37 ± 1.02 -30.33 ± 1.02 -28.41 ± 1.02 -19.93 ± 1.04 -30.02 ± 1.04 - -25.29 ± 1.01 -30.65 ± 1.01 -27.39 ± 1.01 -31.30 ± 1.02 T6 -29.61 ± 1.02 -32.89 ± 1.02 -29.05 ± 1.02 - -31.00 ± 1.04 -30.02 ± 1.02 -25.28 ± 1.01 -29.81 ± 1.01 -25.38 ± 1.01 -31.98 ± 1.02 T7 -33.72 ± 1.02 -34.33 ± 1.02 -28.55 ± 1.02 -17.09 ± 1.04 -31.34 ± 1.04 -30.68 ± 1.02 -27.56 ± 1.01 -31.04 ± 1.01 -27.69 ± 1.01 -30.24 ± 1.02 T8 -28.34 ± 1.02 -30.49 ± 1.02 -27.05 ± 1.02 -20.36 ± 1.04 -29.56 ± 1.04 -27.38 ± 1.02 -24.92 ± 1.01 -28.86 ± 1.01 -24.52 ± 1.01 -27.36 ± 1.02 (B) 15:0 16:0 16:1ω7c 18:0 Bc -24.97 ± 1.01 -29.18 ± 1.01 -28.60 ± 1.01 -28.91 ± 1.02 Co -25.74 ± 1.01 -30.12 ± 1.01 -29.13 ± 1.01 -30.09 ± 1.02 (C) Treatment T1 T2 T3 T4 T5 T6 T7 T8 a15:0 Bc -25.13 ± 1.02 -22.62 ± 1.02 -25.12 ± 1.02 -22.69 ± 1.02 -24.28 ± 1.02 -22.73 ± 1.02 -22.32 ± 1.02 -22.36 ± 1.02 Co -24.78 ± 1.02 -25.17 ± 1.02 -24.26 ± 1.02 -24.00 ± 1.02 -25.81 ± 1.02 -23.81 ± 1.02 -23.34 ± 1.02 -23.05 ± 1.02 i16:0 Bc -25.45 ± 1.02 -25.22 ± 1.01 - - - -25.39 ± 1.01 -27.63 ± 1.01 -25.21 ± 1.01 Co -26.00 ± 1.01 -28.55 ± 1.02 - - - -26.12 ± 1.01 -28.25 ± 1.01 -25.79 ± 1.01