Pathogenic bacteria and mineral N in soils following the land spreading of biogas digestates and fresh manure

Applied Soil Ecology 49 (2011) 18–25

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Pathogenic bacteria and mineral N in soils following the land spreading of biogas digestates and fresh manure M. Goberna a,b,∗ , S.M. Podmirseg a , S. Waldhuber a , B.A. Knapp a , C. García b , H. Insam a a b

University of Innsbruck, Institute of Microbiology, Technikerstrasse 25 d, A-6020 Innsbruck, Austria Centro de Edafología y Biología Aplicada del Segura, Campus Universitario de Espinardo, E-30100 Espinardo, Murcia, Spain

a r t i c l e

i n f o

Article history: Received 12 October 2010 Received in revised form 9 May 2011 Accepted 9 July 2011 Keywords: Anaerobic digestion BIO4GAS Listeria Nitrate leaching Pathogen suppression Salmonella

a b s t r a c t The on-farm production of renewable energy from animal manures has rapidly expanded in central and northern Europe, with thousands of anaerobic reactors. This process has increased the land spreading of biogas digestates, replacing the use of fresh manure as a fertiliser. The environmental benefits and risks of such a change still need to be defined. We hypothesised that applying to the soil anaerobically digested instead of fresh manure might control the release of pathogens but increase that of inorganic N. Pots including ␥- or non-irradiated soils, either control or amended with digestate or manure (80 kg N ha−1 ), were incubated for 0, 1 and 3 months. Escherichia coli, Salmonella and Listeria were cultivated and pathogenicity genes invA and hlyA PCR-amplified. Soil ammonium and nitrate concentrations, and their leaching through the upper soil layer was quantified in 20 cm-depth lysimeters for 100 d. Anaerobic digestion significantly sanitised the manure by completely eliminating cultivable E. coli and Salmonella but not Listeria (1.7 × 104 CFUs g−1 ). Thus, manure increased all microbes and invA gene numbers when applied to soils, while the digestate supplied only hlyA-negative Listeria. Potential pathogens were significantly more abundant in ␥- than in non-irradiated treatments indicating suppression by indigenous soil microbiota. Control levels of all potential pathogens were recovered after 3 months, which could be thus considered a safe delay between land spreading amendments and harvesting. Concentration of nitrates in soil and their movement through the upper layer in soils amended with digestate were doubled compared to the other treatments. Hence, care should be taken that in the field nitrate liberation does not exceed plant demand. © 2011 Elsevier B.V. All rights reserved.

1. Introduction Over 1500 million tonnes of animal manures are produced yearly in EU-27 (Faostat, 2003, in Holm-Nielsen et al., 2009) that are increasingly used in more than 4200 farm-scale anaerobic bioreactors (Tabajdi, 2007). Anaerobic digestion (AD) reduces the volume of wastes while producing biogas, a renewable energy source (Holm-Nielsen et al., 2009; Insam et al., 2010). A co-product of AD is the digestate that can be applied to the soil as an organic amendment (Teglia et al., 2011), contributing to restoring the soil’s organic stock, the recycling of organic matter and the saving of fertilisers. Several features of the digestate have been claimed to make it more appropriate for land spreading than undigested products such as cattle manure (US-EPA, 2005; O’Flaherty et al., 2010). These characteristics include malodour reduction, pathogen control, decreased

∗ Corresponding author at: Centro de Edafología y Biología Aplicada del Segura, Campus Universitario de Espinardo, E-30100 Espinardo, Murcia, Spain. Tel.: +34 968 39 63 80; fax: +34 968 39 62 13. E-mail address: [email protected] (M. Goberna). 0929-1393/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.apsoil.2011.07.007

biochemical oxygen demand, a more balanced nutrient mix and higher nutrient bioavailability, but still no increase in nutrient leaching losses (US-EPA, 2005). However, scientific literature concerning pathogen survival and the leaching of mineral nitrogen is not so straightforward. Pathogenic microbes are not of great concern in liquid and gaseous phases during the anaerobic digestion of wastes (Vinnerås et al., 2006). However, pathogens can survive the process and persist in the digestate (Sahlström et al., 2004; Bagge et al., 2005). The proper sanitation of the end-product depends on the quality of the substrates fed into the reactor, and on the reactor performance, such as previous pasteurisation, digestion temperature, slurry retention time, pH and ammonium concentration, among others (Sahlström, 2003; Ottoson et al., 2008). In the reactor, the numbers of spore-formers, which are commonly found in animal wastes (Snell-Castro et al., 2005), are not reduced (Olsen and Larsen, 1987; Sahlström et al., 2004; Bagge et al., 2005; Goberna et al., 2009). There is also a potential for pathogen regrowth during storage (Sidhu et al., 2001; Pepper et al., 2006), probably due to the non-hygienic conditions of the storage/transporting tanks (Bagge et al., 2005) or to reinoculation by animal vectors

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Table 1 Main characteristics of cattle manure, digestate, soils and ␥ sterilised soils. Manure and digestate data result from a single composite sample. Soil data are given for n = 5 (standard deviations in brackets). Asterisks indicate significant differences between non- and ␥-sterilised soils after t-test analyses (p < 0.05). All data are expressed on a dry weight basis. Parametersa

Manure

Digestate

Soil

␥ Sterilised soil

pH (1:10 soil:CaCl2 ) EC (dS m−1 ) C (g kg−1 ) N (g kg−1 ) C/N NH4 + -N (g kg−1 ) NO3 − -N (g kg−1 ) Dry weight (%) Texture Al (%) B (ppm) Ca (%) Cd (ppm) Cr (ppm) Cu (ppm) Fe (g kg−1 ) K (%) Mg (%) Mn (ppm) Na (%) Ni (ppm) Pb (ppm) P (%) S (%) Zn (ppm)

8.4 3.4 353.5 14.0 25.3 0.87 0.083 11.5 – – – – – – – – – – – – – – – – –

8.6 6.5 320.3 22.8 14.1 12.8 0.34 3.1 – – – – – – – – – – – – – – – – –

7.8 (0.07) * 0.27 (0.01) * 50.0 (6.6) 2.7 (0.6) 19.1 (3.2) 0.002 (0.0005) * 0.025 (0.002) 74.4 (0.3) Loamy-silty 3.19 (0.12) 9.92 (0.88) 2.78 (0.35) 1.05 (0.08) 66.2 (3.36) 63.8 (3.39) 38.7 (2.40) 0.95 (0.05) 2.14 (0.20) 806 (30) 0.08 (0.01) 30.7 (1.1) 66.8 (3.3) 0.19 (0.01) 0.07 (0.01) 124 (4)

7.4 (0.04) 0.29 (0.01) 54.3 (1.6) 2.8 (0.2) 19.2 (0.9) 0.028 (0.001) 0.022 (0.002) 73.9 (0.4) Loamy-silty 3.11 (0.19) 9.89 (0.86) 2.67 (0.21) 1.14 (0.11) 62.7 (3.9) 64.8 (5.60) 36.6 (4.52) 0.92 (0.07) 2.02 (0.09) 808 (65) 0.07 (0.01) 30.6 (1.7) 67.3 (5.2) 0.18 (0.01) 0.07 (0.01) 126 (11)

a

EC, electrical conductivity.

(Zaleski et al., 2005). Pathogens might thus be spread together with the digestate onto agricultural soils. The European Regulations (EC) 1774/2002 and 1069/2009 allow, respectively, the use of manure and digested residues as organic fertilisers and soil improvers (EU, 2002, 2009). These Regulations set the limits of the pathogenic loads admitted in both types of amendments, while stating that further scientific advice is required on such matter (EU, 2002). Mineralisation of the organic N compounds during anaerobic digestion leads to increased levels of the soluble forms of inorganic N, mainly ammonia (Möller and Stinner, 2009). During application to the land, up to 15% of the total applied N is emitted to the atmosphere as ammonia (Sommer and Hutchings, 2001; Matsunaka et al., 2006; Möller and Stinner, 2009; TerhoevenUrselmans et al., 2009). However, this percentage varies depending on the type of storage and land spreading of the slurry, as well as the environmental conditions during the application (Sommer and Hutchings, 2001; Sandars et al., 2003; Holm-Nielsen et al., 2009). These and other factors (e.g. infiltration of ammonium into the soil) also influence the gaseous N emissions after application (Sommer and Hutchings, 2001). Once in the soil matrix, ammonium is mostly adsorbed to clays and organic matter through its positive charge, transformed into microbial biomass N or nitrified, i.e. turned into nitrite and nitrate or further into molecular N (Amlinger et al., 2003). Increased inorganic N surface run-off or leaching through the soil profile after application of organic and inorganic fertilisers occurs mostly in the form of nitrate (Insam and Merschak, 1997; Matsunaka et al., 2006), due to its high solubility and negative charge that decreases its adsorption to soil colloids. Based on the reduction of organic N during the process of anaerobic digestion, Ørtenblad (2002) predicted that 8% more N would be lost by leaching after land spreading cattle excreta than biogas digestates. However, recent research does not support such calculations (Matsunaka et al., 2006; Möller and Stinner, 2009) most likely because the nitrogen contained in the manure is more stable (e.g. lignocellulosic materials) compared to the larger amounts of mineral N forms present in the anaerobically digested products.

Therefore, more experimental data need to be gathered in order to better define the benefits and risks of land spreading biogas digestates to the environment. We simulated, at a microcosm level, a single application event of either cattle manure or digested manure into both ␥-sterilised and non-sterilised soils. Our objectives were to (i) analyse the survival of selected pathogens, (ii) judge the role of the indigenous microbiota in outcompeting potential pathogens and (iii) quantify soil ammonium and nitrate concentrations and leaching through the topsoil. We hypothesised that using digested instead of undigested manure might result in reduced pathogen survival, due to certain sanitation during anaerobic digestion, but increased N percolation owing to higher ammonium and nitrate concentrations. We expected that pathogens growing in the amendments (if any) would grow better in the sterilised soils due to the absence of competitors.

2. Materials and methods 2.1. Soils, cattle manure and biogas digestate Soils, cattle manure and biogas digestate were collected from the agricultural school in Rotholz (Tirol, Austria). In this area, cattle manure used to be applied to arable soils three times per year at a maximum rate of 80 kg N ha−1 at each single application. Nowadays, an equivalent amount of digestate, which is produced at the on-farm biogas reactor, is applied instead. This is a BIO4GAS® plant (Wett and Insam, 2010), which was started up in February 2008 by completely filling it up with a mixture of solid and liquid cattle excreta (plus bedding straw) and progressively heating it up to 37 ◦ C. From the plant startup until our sampling, the reactor was fed continuously with 5 m3 cattle excreta d−1 , corresponding to a hydraulic retention time of 60 d. A total of 150 kg topsoil (20 cm depth) was collected as a composite sample from a 5 m × 10 m agricultural plot on May 28, 2008. Soil was sieved (<4 mm) and visible roots and animals removed. This was divided into two 70 kg subsamples: one was stored at 4 ◦ C

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Table 2 PCR amplification of the invA and hlyA genes with primers specific for Salmonella and Listeria monocytogenes, respectively, in all treatments. Symbols are as follows: +, indicates positive amplification of at least one replicate out of five at any incubation time; −, indicates no amplification from any of the replicated samples or incubation times. invA

hlyA

Experimental treatments Control soil (C) Soil + Digestate (D) Soil + Manure (M) Sterilised control soil (␥C) Sterilised soil + Digestate (␥D) Sterilised soil + Manure (␥M)

– – – – – +

– – – – – –

Amplification controls Positive controla C spiked with positive controla

+ NAb

+ +

a Positive controls were Salmonella enterica (DSMZ 10062) and Listeria monocytogenes (DSMZ 15675). b Not analysed.

and the other one was ␥-sterilised at 25 kGy using Co 60 as source for radiation (Mediscan, Kremsmünster, Austria). Cattle manure and digestate were collected on June 18, 2008. Cattle manure resulted from mixing solid excreta (plus bedding straw) and liquid excreta (1:1, w/v), from stabled cows. Digestate was sampled from the last sampling port, i.e. the effluent material after a 60 d digestion, in the BIO4GAS® plant located nearby the stable. Freshly collected cattle manure and digestate were used for chemical analyses (Table 1). Subsamples of both materials were air-dried for 24 h, and then dried at 45 ◦ C for 24 h prior to C and N analyses (Leco TruSpec Macro CHN). Soil and ␥ sterilised soil were air-dried prior to chemical analyses (Table 1). 2.2. Experimental set-up and sampling Six experimental treatments resulted from the combination of three types of amendments (none, cattle manure and digestate) to both ␥-sterilised and non-sterilised soils (see Table 2 for treatments and abbreviations used hereafter). Cattle manure or biogas digestate were mixed with soil heaps by turning at a rate of 40 mg N kg−1 soil (dry weight). This amount is equivalent to 80 kg N ha−1 , considering the soil bulk density of 1 g cm−3 and a plough depth of 20 cm, and was selected to simulate a single application event. Soil water content was adjusted to 50% water-holding capacity (100% WHC = 0.75 mL H2 O g−1 soil) in all treatments with autoclaved distilled water. The experiment was set up in autoclaved columns (11 cm diameter, 20 cm depth). A total of 90 columns (2 sterilisation levels × 3 amendment levels × 3 incubation times × 5 replicates) were set up containing 2000 g soil each (wet weight) and arranged in a completely randomised design. After an equilibration period of 4 d at 4 ◦ C, 30 samples were collected (incubation time 0 months). The other 60 samples were incubated in a growth chamber at 20 ◦ C. This is approximately the average temperature of the hottest and wettest month in the area, which would be the most beneficial conditions for pathogen survival (Venglovsky et al., 2009). Thirty of the incubated samples were collected after 1 month and the remainder 30 after 3 months. Three extra columns containing control soil were incubated throughout the whole period for humidity control, which was adjusted as needed in all columns. 2.3. Pathogen cultivation Cattle manure, digestate and soil suspensions corresponding to all six treatments and replicates (1:10 sample:0.95% NaCl) were

shaken at 200 rpm for 15 min. After letting them settle for 1 h, 1/10 serial dilutions from the supernatant down to 10−5 were prepared in 0.95% NaCl. A volume of 30 ␮L was plated on selective agar plates and incubated at 37 ◦ C. E. coli was cultivated on TBX Chromocult Agar (Merck 1.16122), Salmonella spp. on Salmonella Shigella Agar (OXOID CM0099) and Listeria spp. on PALCAM-Listeria-Selektivagar (Merck 1.11755 and supplement 1.12122). All plates were prepared within 24 h after sampling. Colony forming units (CFUs) were counted after 48 h. Plates with no CFUs were re-checked for growth for 2 more weeks. Colourless colonies with black central spots were assumed to be Salmonella spp. and black colonies either L. monocytogenes or L. innocua, according to the respective instruction manuals. 2.4. Molecular detection of pathogens Soil DNA was extracted using the PowerSoil DNA isolation kit (MO BIO Laboratories, Carlsbad, CA) within 24 h after sampling. The invA gene coding for invasine, a protein largely determining the invasive abilities of several pathogens, was amplified using the Salmonella specific primers INVAF (5 -CGGTGGTTTTAAGCGTACTCTT-3 ) and INVAR (5 CGAATATGCTCCACAAGGTTA-3 ) by Fratamico and Strobaugh (1998). PCR amplifications were performed in a Flexcycler (Analytik Jena, Germany) in 25 ␮l volumes, with each reaction containing a final concentration of 1× reaction buffer [16 mM (NH4 )2 SO4 , 67 mM Tris–HCl pH 8.8, 1.5 mM MgCl2 , 0.01% Tween 20], 200 ␮M each dNTP, 0.2 ␮M each primer, 1 mM MgCl2 , 0.1× enhancer (Peqlab, Germany), 0.625 U BioThermTM DNA polymerase (GeneCraft, Germany), and sterile water. A volume of 2 ␮l DNA was directly applied to the reaction mix. Thermal cycling started with 2 min at 94 ◦ C, followed by 38 cycles consisting of 20 s at 94 ◦ C, 60 s at 53 ◦ C and 60 s at 72 ◦ C, and a final step at 72 ◦ C for 10 min. A pure culture of Salmonella enterica (DSMZ 10062; Deutsche Sammlung von Mikroorganismen und Zellkulturen, Germany) was used as a positive control. The hlyA gene, encoding the virulence factor listeriolysin O, which is essential for the virulence of L. monocytogenes, was amplified with the specific forward (5 -TGCAAGTCCTAAGACGCCA3 ) and reverse primers (5 -CACTGCATCTCCGTGGTATACTAA-3 ) by Nogva et al. (2000). PCR reaction mix was prepared as above and thermal cycling included 5 min at 95 ◦ C, 40 cycles consisting of 20 s at 95 ◦ C, 20 s at 58 ◦ C and 60 s at 72 ◦ C, and a final step at 72 ◦ C for 10 min. L. monocytogenes (DSMZ 15675) was used as a positive control. PCR products (6 ␮L) obtained from the amplification of all soil samples were loaded on 3% agarose gels in 1× Tris-acetateEDTA (TAE) buffer. Electrophoresis was run at 50 V for 60 min. The GeneRulerTM Express DNA ladder (Fermentas International Inc., Canada) was used as a molecular weight marker. Gels were stained with ethidium bromide (0.1%, v/v), inspected under UV light and photographed with a digital camera (Power Shot A 640, Canon, China). 2.5. Ammonium and nitrate leaching The fifteen columns containing non-sterile soils, both control and soils amended with cattle manure or digestate, which had been incubated for 1 month at 20 ◦ C were used for the leaching experiment. Prior to the start of the leaching experiment, soil cores (1 cm diameter × 2 cm depth) were taken for analysing NH4 + N and NO3 − -N concentrations. Columns were then watered with distilled water up to 100% WHC and with 100 mL extra to force leaching. Watering was automated using a dispensing pump BKV MS/CA8-6 (Ismatec, Switzerland) with a flux rate of 5 mL h−1 . For an incubation period of 100 d, the columns were kept at 20 ◦ C.

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This temperature approximately corresponds to the average temperature of the hottest and wettest month in the region, which would maximise N mobility through the profile (Gundersen and Rasmussen, 1995). Leaching was repeated weekly. Ammonium-N (NH4 + -N) and nitrate-N (NO3 − -N) in the leachates were determined as described by Kandeler (1993a,b). Five mL leachate was mixed with 2.5 mL of a freshly prepared solution consisting of 0.1 M sodium hydroxide, 0.35 M sodium nitroprusside dihydrate and 4.02 mM sodium salicylate. One mL of 0.1% (w/v) dichloroisocyanuric acid sodium salt dihydrate was added to the mix. Standard solutions containing 0.0, 1.0, 1.5, 2.0, 2.5 ␮g N mL−1 were prepared and treated the same way as the samples. After 30 min incubation, NH4 + -N was determined spectrophotometrically at 660 nm. NO3 − -N was determined after adding 10 mL of 1/100 diluted leachate and 0.4 mL 10% sulphuric acid to two test tubes per sample. Two copper-sulphate covered zinc granules were added to one of the tubes. Standard solutions containing 0.0, 0.5, 1.0 and 1.5 ␮g N mL−1 were treated as the samples. After an overnight incubation, NO3 − -N was determined spectrophotometrically at 210 nm. Absorbance values of tubes containing zinc granules were subtracted from their corresponding granule-free tubes. 2.6. Statistical analyses Non-parametric tests for several independent samples (Kruskal–Wallis test) were performed to test for differences in cultivable pathogens (log-transformed data) among treatments and incubation times. The ability of allochthonous microbes to grow on ␥-sterilised or non-sterilised soils was evaluated by comparing the CFUs after subtracting the numbers in the amendment from the corresponding control. Pair-wise comparisons between treatments were performed using the Mann–Whitney test. Univariate analysis of variance (ANOVA) was used to test for differences in NH4 + -N and NO3 − -N contents in soils before the percolation experiment, as well as in NH4 + -N in the leachates on the first leaching event. Repeated measures ANOVA was used to assess the effects of the inter-subject factor “amendment” on nitrate leaching using “incubation time” as intra-subject factor, and Tukey’s test for post hoc mean separation. All statistical analyses were performed with SPSS 17.0. 3. Results 3.1. Pathogen counts and detection of pathogenicity genes Cultivable E. coli, Salmonella and Listeria were more abundant in the cattle manure than in the digestate (Fig. 1). These differences were significant for Salmonella (U < 0.0001, p = 0.005) and Listeria (U < 0.0001, p = 0.009). E. coli could not be isolated from the digestate, but differences with fresh manure were not significant (U = 7.50, p = 0.14) due to the high variance in the number of CFUs (note that the y-axis in Fig. 1 has a logarithmic scale). E. coli and Salmonella showed similar patterns upon soil application (Fig. 1). CFUs were detected only in cattle manure and treatments M, ␥M and ␥D (only for Salmonella). E. coli CFUs in cattle manure were not significantly different from those in M or ␥M, if detected (U ≥ 7.5, p ≥ 0.1), whereas significantly more Salmonella were counted in the manure compared to M and ␥M (U ≤ 0.001, p ≤ 0.008). Treatment ␥M had significantly more Salmonella than M (U = 61.0, p = 0.01; after subtracting the numbers in the respective controls). Listeria had ≥240 CFUs g−1 soil in all the non-sterile treatments (Fig. 1). The control soil (C) had an average of 1750 CFU g−1 at the start of the experiment. Amending with manure (M) or digestate (D) induced significantly higher Listeria CFUs compared with the

Fig. 1. Escherichia coli, Salmonella sp. and Listeria sp. in the amendments (digestate and manure) and in the amended soils (soil + digestate, soil + manure, sterilised soil + digestate and sterilised soil + manure) compared to their controls (control soil or sterilised control soil) at three incubation times (0, 1 and 3 months). Bars indicate standard errors (n = 5). Note that the comparison between sterilised and non-sterilised treatments was made after subtracting the CFUs in the control soils (black contour) from the CFUs in the amended soils.

control soil after 1 month incubation (U = 0.5, p = 0.008), but after 3 months both treatments were similar to the control (U ≥ 9.0, p ≥ 0.45). Differences between M and D were not significant at any time (U = 106.5, p = 0.80), whereas ␥D had significantly more CFUs than ␥M (U = 29.0, p < 0.001). Treatments ␥M and ␥D had significantly higher Listeria counts than ␥C at all incubation times (U ≥ 8.5, p > 0.001). The numbers of Listeria added with the amendments (CFUs amended soil–CFUs control soil) were significantly more abundant in ␥M and ␥D compared to M and D, respectively (U ≥ 12.0, p ≤ 0.002). The invA and hlyA genes of the pure cultures used as positive controls were PCR amplified (Table 2) and gave products of the expected length (796 and 113 bp, respectively). InvA specific for Salmonella was only detected in one replicate (out of five) of treatment ␥M (Table 2). L. monocytogenes was not detected in any soil sample (Table 2). Spiking a soil DNA extract with the L. monocytogenes pure culture (1/10 ratio) resulted in positive amplification (Table 2), indicating negligible presence of PCR inhibitors in the soil matrix. 3.2. Ammonium and nitrate leaching NH4 + -N and NO3 − -N concentrations in the digestate were 15and 4-fold those of cattle manure, respectively (Table 1). However, NH4 + -N concentrations in soils immediately before leaching was induced were not significantly different among treatments (F = 0.89, p = 0.44), whereas those of NO3 − -N were significantly higher in soils amended with digestate (F = 6.17, p = 0.014; Fig. 2). During the 100 d percolation experiment, similar leachate volumes were recovered from all soil columns (ca. 100 mL per leaching event). Measurable amounts of NH4 + -N were obtained only in the

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M. Goberna et al. / Applied Soil Ecology 49 (2011) 18–25 Table 3 Number of E. coli and Salmonella (CFU g−1 dry weight) in cattle manure and digestate samples, and thresholds established by the European Regulation (EC) No 1774/2002 (EU, 2002). Sample nr.

E. coli

Salmonella

Manure

Fig. 2. Total N (g kg−1 ), NO3 − -N and NH4 + -N (mg kg−1 ) concentrations immediately before starting of percolation experiment in the control soils and soils amended with digestate or manure. Bars indicate standard errors (n = 5 in all cases but n = 4 for soils amended with manure). Different letters indicate significant differences among treatments (p < 0.05) for each variable.

first leaching event and did not differ among treatments (F = 1.32, p = 0.31). The leachate from the control soil contained (mean ± SE) 0.20 ± 0.06 ␮g NH4 + -N mL−1 , and those from the soils amended with digestate or manure had 0.25 ± 0.02 and 0.29 ± 0.04 ␮g NH4 + N mL−1 , respectively. NO3 − -N content was significantly higher in the leachates from D compared to C and M (F = 6.50, p = 0.016); C and M could not be statistically differentiated (Fig. 3). The significant interaction among inter- and intra-subject factors (F = 4.68, p = 0.005) allowed using one-way ANOVA for treatment comparison at each sampling date. This indicated that nitrate leaching was significantly higher in D compared to C and M in measurements from days 6 to 34 (asterisks in Fig. 3). No significant differences among treatments were detected on day 1 and from day 41 on. 4. Discussion 4.1. Amending with digestate released less potential pathogens Salmonella spp., Listeria spp. and E. coli are common bacteria in livestock wastes which can cause gastroenteritis-type diseases (Mawdsley et al., 1995). All of them were isolated from cattle manure, i.e. mixed urine, faeces and bedding material, collected from stabled animals in this study. However, no culturable forms

Fig. 3. Nitrate leaching from control soils and soils amended with digestate or manure during a 100 d incubation in 20 cm-depth lysimeters. Bars indicate standard errors (n = 5 in all cases but n = 4 for soils amended with manure). Asterisks indicate significant differences (p < 0.05) among treatments for each leaching event.

Digestate

Manure

Digestate

1 2 3 4 5

1.4 × 10 0 0 0 5.5 × 103

0 0 0 0 0

2.6 × 10 2.8 × 105 2.8 × 105 1.6 × 104 2.8 × 105

0 0 0 0 0

Legislated valuesa m M c

0 103 5

103 5 × 103 1

0 0 0

0 0 0

Meets legislation

No

Yes

No

Yes

3

4

a

Standard values according to EU (2002). m, threshold value for the number of bacteria (satisfactory if all samples contain ≤m); M, maximum value for the number of bacteria (unsatisfactory if one or more samples ≥M); m and M are given for 1 g (E. coli) or 25 g (Salmonella), and measured in 5 samples; c, number of samples (out of a total of 5) the bacterial count of which may be between m and M (acceptable if the bacterial count of the other samples is ≤m).

of Salmonella or E. coli were detected in the end-product after the anaerobic digestion in the BIO4GAS® plant located near the stable. Similarly, fewer Listeria were detected in the digestate compared to the manure. Hence, certain sanitation of the animal wastes was achieved because of anaerobic digestion, as has been reported previously (Bagge et al., 2005). In mesophilic digesters, the decimation time or time required to reduce by 90% the viable counts of a microbial population can be counted in days, compared to weeks or months when material is simply stored (Sahlström, 2003). For instance, Olsen and Larsen (1987) calculated that populations of Salmonella sp. and E. coli had decimation times shorter than 3 days in laboratory- and full-scale mesophilic reactors. Thus, the long sludge retention time (60 days) in the BIO4GAS® reactor might underlie its efficient killing of the studied pathogens. On the other hand, the large ammonia contents in this manure-digesting reactor could have further promoted the sanitation of the product (Ottoson et al., 2008). However, the survival of Listeria was remarkable (1.7 × 104 CFU g−1 digestate) even after the 60-day anaerobic digestion. Kearney et al. (1993) suggested that nutrient availability could be a major cause of differential survival of pathogenic microorganisms in biogas digesters. In a continuously fed reactor, nutrient supply is determined by the rate at which the reactor is fed (Sahlström, 2003). This indicates that feeding a mesophilic BIO4GAS® reactor with 5 m3 cattle excreta d−1 provides enough nutrients so as to maintain a large population of Listeria. Therefore, under such conditions, there exists a risk of spreading potentially pathogenic microbes onto agricultural fields while fertilizing with anaerobically digested cattle excreta, as others have warned for several types of biowastes (Bagge et al., 2005; Sahlström, 2003; Sahlström et al., 2004). As these authors recommended, including an initial pasteurisation step or increasing the temperature of anaerobic digestion to 55 ◦ C could further sanitise the end-product. Even though, in our study, the anaerobic treatment transformed the cattle manure, which did not reach the European legislated standards on pathogenic loads in animal by-products not intended for human consumption, into a digested product meeting the current regulations (EU, 2002; Table 3). The agricultural soils in our study site did not contain detectable levels of culturable Salmonella or E. coli. However, over 1.7 × 103 CFUs g−1 Listeria were counted in the control soil. These organisms could have persisted in the soil after the previous field applications of animal wastes, which are generally land-spread in

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two or three loads per year. It is known that E. coli does not survive longer than some days or weeks in the soil, although as reviewed by Unc and Goss (2004) some authors detected survival periods of up to 30–68 d under saturated conditions or after application of pig manure. However, some species of Salmonella have been found to persist up to 110 d (Mawdsley et al., 1995) and Listeria up to 128 d (Venglovsky et al., 2009). The same agricultural soils amended with a dose equivalent to 80 kg N ha−1 of digestate did not contain cultivable forms of Salmonella or E. coli, whereas both were present in soils amended with manure. Also, the invA gene of Salmonella was detected in sterilised soils amended with manure, indicating the land-spreading of pathogenic strains supplied with the untreated wastes. Similarly, spreading manure resulted in E. coli O157 and Salmonella survival for up to 1 month under field conditions (Nicholson et al., 2005). However, Salmonella was present at very low levels in soils amended with anaerobically digested biosolids, becoming undetectable 2 weeks after the application of the amendment (Zaleski et al., 2005). In our study, soils amended with either cattle manure or digestate had significantly higher numbers of Listeria than the control after 1 month incubation. This demonstrates the survival in the environment for up to 1 month of potential pathogens contained in either type of amendment. Similarly, Listeria spread with manure survived for over a month after land-spreading under field conditions (Nicholson et al., 2005). The number of Listeria thriving in the soils in our experiment were extremely high, considering that as few as 1 CFU Listeria g−1 soil can be transferred to vegetable leaves by splashing of rainfall or irrigation (Girardin et al., 2005). However, PCR detection of the hlyA gene with primers specific for L. monocytogenes gave negative results. This indicates that cultivable forms of Listeria in the soils studied could correspond to L. innocua, a common and harmless soil inhabitant which shares morphological, biochemical and molecular features with L. monocytogenes (Liu et al., 2003). Still, the possibility that viable but non-culturable forms of unknown pathogenicity do exist should not be excluded (Mawdsley et al., 1995). After 3 months, the numbers of cultivable potential pathogens in the amended soils did not exceed those in the control soils and all PCR assays gave negative results, despite incubation under constant mild temperatures and humid conditions increasing pathogen survival (Venglovsky et al., 2009). This suggests that 90 days could be a reasonable period of delay between land-spreading organic amendments and crop harvesting. This is in keeping with previous reports, although pathogen survival in soils depends on a number of factors, including the environmental conditions, the quality of waste applied and the microbial species (Lepeuple et al., 2004; Girardin et al., 2005). Finally, it should be considered that even if the application of the digestate to the soil released fewer potential pathogens, any microbe supplied with the digestate, with its low solid content, could have higher mobility and more easily reach deeper soil layers (Unc and Goss, 2004). 4.2. Indigenous soil microbiota suppressed potential pathogens The survival in soils of potential pathogens supplied with the amendments was evaluated by comparing previously ␥irradiated soils with non-irradiated soils, after subtracting the CFUs in amended soils from those in the control. Potential pathogens supplied with the amendments, when present, were more abundant in previously sterilised soils (although differences were not significant for E. coli due to high data variance). This demonstrates that potential pathogens proliferated better in soils lacking an autochthonous microbiota, that is, under conditions of relaxed competition and large niche availability. Microbial competition between indigenous soil microbes and pathogens has been proven in laboratory experiments using co-cultures. For exam-

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ple, Buchanan and Bagi (1999) demonstrated that Pseudomonas fluorescens is able to limit the growth of L. mononocytogenes, the intensity of the interaction depending on the environmental conditions (mostly temperature and sodium chloride concentration). The suppressive capacity of soils and composts against pathogens has been also reported both for vegetative cells and spores by comparing sterilised and non-sterilised material. S. typhimurium and S. newport grew well in irradiated and died in non-sterilised composts suggesting suppression by the indigenous compost microbiota (Hussong et al., 1985; Sidhu et al., 2001). Petras and Casida (1985) inoculated spores of Bacillus thuringiensis in the soil at field and laboratory trials. Spores survived but did not have the ability to germinate, what led the authors to allude to sporostatic effects mediated by microorganisms, secondary metabolites or extracellular enzymes. More recent research has evidenced pathogen suppression by soil microbes through the production of antibiotics, the deprivation of nutrients mediated by chelators and the interference with pathogenicity factors (reviewed by Haas and Défago, 2005). It should be considered that ␥-radiation can increase nutrient availability (Eno and Popenoe, 1964) and alter the structure of organic matter most probably through microbe lysis inducing an increased proportion of the light fractions (Berns et al., 2008). We did not find differences between ␥- and non-irradiated control soils as regards organic C and N, total macro- and micronutrients or NO3 − -N, but ␥-sterilisation increased NH4 + -N concentration 14-fold (Table 2). Therefore, exogenous microbiota applied to ␥irradiated soils could have partly benefited from this artificial increase in the availability of mineral N. Despite this artefact, ␥radiation is considered to be the most effective method for soil sterilisation having the least impact on soil properties (Trevors, 1996).

4.3. Digestate increased soil nitrate concentration and percolation The process of anaerobic digestion largely mineralised the organic N contained in the fresh manure, as expected. The proportion of mineral N (NH4 + -N + NO3 − -N) to total N increased from 6.8% in manure to 57.5% in biogas digestate. This highlights the adequacy of digestates as fertilisers since their higher mineral N contents both reduce the needs for supplemental mineral N fertilisation and the potential for residual mineral N after crop harvesting. However, these results also confirm that land spreading digestates bears higher risk of inorganic N release to the environment. We compared the N leaching potential through the topsoil of fresh and anaerobically digested cattle manure by forcing water percolation in 20 cm-depth lysimeters after applying the amendments to the soil at an equivalent N concentration of 40 mg N per kg soil. This allows a straightforward comparison between both amendment types under controlled conditions, while reducing the variability of results obtained under field conditions due to the multiple interacting factors (e.g. Möller and Stinner, 2009). However, the figures derived from such an experiment cannot be taken as representative of the concentrations that would be drained to the surface waters or leached to the groundwater due to the following reasons. First, the production of a leachate requires working under soil saturation conditions, thus maximising the mobility of the soluble N forms through the profile (Gundersen and Rasmussen, 1995). However, water saturation might also increase denitrification, that is, the microbial consumption of NO3 − -N (Smith and Tiedje, 1979). Second, the use of small lysimeters can also result in the overestimation of the concentrations of the N leached, since the largest amounts of NO3 − -N after application of amendments are found in the upper 60 cm layer (Amlinger et al.,

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2003). Finally, the absence of crops, which uptake mineral N, could yield a similar result. Ammonium was similar in amended and unamended soils immediately before the leaching experiment, that is, after 1 month of incubation at 20 ◦ C and 50% WHC in the absence of water percolation. This suggests that the high NH4 + -N contents present in the organic amendments were lost from the system, most likely through nitrification and, to some extent, through volatilisation during sample handling (Amlinger et al., 2003). Volatilisation during the incubation might have been almost negligible since the amendments were mixed with the soil instead of being spread on the surface (Sommer and Hutchings, 2001). Detectable ammonium leaching occurred only during the first percolation event and was not significantly different between amended and unamended soils. This was not unexpected since Insam and Merschak (1997) also found negligible levels of NH4 + -N in the leachates collected from soil lysimeters amended with several types of fertilisers. Complete retention of added NH4 + -N in the upper soil layers has been also observed in field experiments (Gundersen and Rasmussen, 1995), what can be attributed to adsorption by soil colloids, immobilisation by microbes and transformation of ammonium to nitrate (Amlinger et al., 2003). Significant NH4 + -N leaching has been reported in soils chronically receiving high NH4 + inputs, as discussed by Gundersen and Rasmussen (1995). Soil NO3 − -N concentration just before the leaching experiment was significantly higher in soils amended with digestate compared to the other treatments. Also, the 100 d accumulated nitrate percolation from soils amended with digestate (mean ± SE, 1634 ± 262 ␮g N mL−1 leachate) doubled that of soils amended with manure (829 ± 173 ␮g N mL−1 ) which did not significantly differ from the control (847 ± 132 ␮g N mL−1 ). This implies that 23% of the total N contained in the soil (natural + added) was leached in the form of nitrates during the whole period in soils amended with manure, whereas 45% was lost from soils amended with biogas digestates. The excess nitrate might have come mostly from the nitrification of the large amounts of ammonium in the digestate. This contrasts with the stability of the N forms immobilised in the solid organic fraction of cattle manure. Our results are in agreement with those from Matsunaka et al. (2006), who quantified N leaching at 1.7 m depth from grass cultivated plots amended with anaerobically digested cattle slurry or a mineral fertiliser. These authors found that N leaching was the second most important source of N leakage to the environment, after NH3 volatilisation. The amount of mineral N ranged from 80 to 96% of the soluble N in the leachate depending on the season and application rate. However, this disagrees with the findings by Möller and Stinner (2009), who measured the soil mineral N content from the surface to 90 cm depth after applying fresh farmyard manure, liquid cattle slurry, digested slurry, and digested slurry plus crop residues on two seasons to several crop rotations. In most cases, these authors did not find significant differences in the inorganic N contents in soils amended with digested cattle slurry compared to soils amended with undigested cattle slurry or fresh farmyard manure. However, for particular crop rotations and depending on the residues used for digestion they found either significantly increased or reduced N leaching in the former soils compared to the latter. Differences in the digestion process, the time, form and rate of application to the field, and the chemical properties of the residues applied might underlie the conflicting results found in the literature. In our experiment, significantly higher nitrate losses from soils amended with digestate occurred in five out of the fifteen leaching events analysed. The first event did not produce significantly greater leaching in the soil amended with digestate, indicating a phase of reduced NO3 − -N mobility through the profile. This was probably due to the lower water availability (initial incubation at 50% WHC compared to subsequent incubation at 100% WHC) and

thus water percolation, which largely determines nitrate leaching (Gundersen and Rasmussen, 1995). From the sixth raining event on, no significant leaching occurred as compared to the control, likely due to exhaustion of excess mineral N supplied. Up to that time point (day 34), the amount of nitrate leached was 38% (manure) to 52% (digestate) of the total mineral N leached during the 100 d incubation. This contrasts with the results by Lim et al. (2010), who found that 50–94% of mineral N was leached during the first week after applying manure-based composts to several soil types in a 19 week column experiment. As opposed to us, these authors used air-dried samples, what could have caused the rapid flush of N (Lim et al., 2010). Our results suggest that the use as a fertiliser of digested, instead of fresh, cattle manure releases comparatively more nitrates to the environment. If not taken up by plants, nitrates could be either drained to surface waters, leached to groundwaters or denitrified into gaseous forms and emitted to the atmosphere. Thus, land spreading of digestates should accurately match crop N demand. Pre-treatment options for the digestates should also be considered. An option would be composting of the digestates together with a bulking agent and thus reduce the immediately available nutrients (Insam and Merschak, 1997; Miller et al., 2008). This strategy has been also suggested as a means of stabilising the residual organic matter (Teglia et al., 2011) and reducing the phytotoxicity exhibited by biogas digestates compared to mature composts (Fuchs et al., 2008). 5. Conclusions Mesophilic anaerobic digestion of cattle manure in a BIO4GAS® reactor sanitised the animal wastes efficiently, with a complete destruction of E. coli and Salmonella sp. and a significant reduction in the viable counts of Listeria sp. Therefore, fewer potential pathogens survived in soils amended with digestate compared to those amended with manure. After application to the soil, the indigenous microbiota suppressed the proliferation of pathogens, and 3 months after the application of either type of amendment control levels had been recovered. On the other hand, nitrate percolation increased in the short-term when anaerobically digested instead of fresh cattle manure was used in a lysimeter experiment without plants. Laboratory and field trials varying the application rate and form of digestates in the presence of crops, and using composted digestates, could help optimising the use of biogas digestates for agricultural applications. Acknowledgements Financial support was provided by the EU Marie Curie Programme (MEIF-CT-2006-041034) to MG and HI. The authors thank B. Stojanovic for technical assistance, F. Schinner for the dispensing pump and growth chamber and the staff in the agricultural school in Rotholz for their kind cooperation. The authors acknowledge the critical revision of this manuscript by four anonymous reviewers. References Amlinger, F., Götz, B., Dreher, P., Geszti, J., Weissteiner, C., 2003. Nitrogen in biowaste and yard waste compost: dynamics of mobilisation and availability—a review. Eur. J. Soil Biol. 39, 107–110. Bagge, E., Sahlström, L., Albihn, A., 2005. The effect of hygienic treatment on the microbial flora of biowaste at biogas plants. Water Res. 39, 4879–4880. Berns, A.E., Philipp, H., Narres, H.D., Burauel, P., Vereecken, H., Tappe, W., 2008. Effect of gamma-sterilization and autoclaving on soil organic matter structure as studied by solid state NMR UV and fluorescence spectroscopy. Eur. J. Soil Sci. 59, 540–550. Buchanan, R.L., Bagi, L.K., 1999. Microbial competition: effect of Pseudomonas fluorescens on the growth of Listeria monocytogenes. Food Microbiol. 16, 523–530.

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