Recycling of sewage sludge and household compost to arable land: fate and effects of organic contaminants, and impact on soil fertility

Soil & Tillage Research 72 (2003) 139–152

Recycling of sewage sludge and household compost to arable land: fate and effects of organic contaminants, and impact on soil fertility S.O. Petersen a,∗ , K. Henriksen b , G.K. Mortensen c , P.H. Krogh d , K.K. Brandt e , J. Sørensen e , T. Madsen f , J. Petersen a , C. Grøn c,f,1 a

d

Department of Agroecology, Danish Institute of Agricultural Sciences, Tjele, Denmark b Department of Environmental Engineering, Aalborg University, Aalborg, Denmark c Risø National Laboratory, Plant Research Department, Roskilde, Denmark Department of Terrestrial Ecology, National Environmental Research Institute, Silkeborg, Denmark e Department of Ecology, Royal Veterinary and Agricultural University, Copenhagen, Denmark f DHI Water and Environment, Hørsholm, Denmark Received 21 November 2002; accepted 3 March 2003

Abstract Effective use of organic wastes for agricultural production requires that risks and benefits be documented. Two types of sewage sludge, household compost and solid pig manure were studied under field and greenhouse conditions to describe their fertilizer value and effects on soil properties and soil biota, the fate of selected organic contaminants, and their potential for plant uptake. A 3-year field trial on two soil types showed no adverse effects of waste amendment on crop growth, and a significant fertilizer value of one sludge type. Accumulation of N and Pi was indicated, as well as some stimulation of biological activity and micro-arthropod populations, but these effects differed between soil types. There was no detectable accumulation of polycyclic aromatic hydrocarbons (PAH), di(2-ethylhexyl)phthalate (DEHP), nonylphenol and ethoxylates (NP + NPE) or linear alkylbenzene sulfonates (LAS) after three repeated waste applications, and no plant uptake was suggested by analysis of the third crop. A plot experiment with banded sludge was conducted to examine sludge turnover and toxicity in detail. Less than 5% of NP or LAS applied in organic wastes was recovered after 6 months, and less than 6% of DEHP applied was recovered after 12 months. Potential ammonium oxidation (PAO) at 0–1 cm distance from the banded sludge was stimulated despite toxic concentrations in the sludge, which suggested that contaminants were degraded inside sludge particles. Phospholipid fatty acid (PLFA) profiles suggested a gradual shift in the composition of the microbial community within sludge, partly due to a depletion of degradable substrates. A pot experiment with sludge-amended soil and soil spiked with contaminants showed no plant uptake of NP, DEHP or LAS. Degradation of LAS and NP added in sludge was delayed and the degradation of DEHP was faster than when the contaminants were added directly to the soil. In conclusion, adverse effects of organic waste application on soil or crop were not found in this study, and for some waste products positive effects were observed. © 2003 Elsevier Science B.V. All rights reserved. Keywords: Fertilizer value; Collembola; Mites; Barley; Oat; Rape; PLFA; Nitrification; Plasticizer; Surfactant

∗ Corresponding author. Tel.: +45-8999-1723; fax: +45-8999-1619. E-mail address: [email protected] (S.O. Petersen). 1 Present address: Eurofins A/S, Hørsholm, Denmark.

0167-1987/03/$ – see front matter © 2003 Elsevier Science B.V. All rights reserved. doi:10.1016/S0167-1987(03)00084-9

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1. Introduction

2. Materials and methods

For agriculture, organic waste materials represent an inexpensive nutrient source and soil conditioner. Recycling of sewage sludge is already common practice; according to Webber et al. (1996) the recycling of sewage sludge was 30% in Canada and 40% in the UK. In Western Europe, current trends in waste management policies favor land application as opposed to land fill deposition or incineration (Council Directive, 1999/31/EC). Efficient use, however, requires that risks and benefits of different categories of organic wastes are thoroughly documented. Positive effects of organic waste application on soil properties have been documented for soil structure, bulk density, water retention characteristics, CEC, and soil biological activities (Levi-Minzi et al., 1985; Dar, 1997; Elsgaard et al., 2001b). The nutrient value of organic wastes is considered to be moderate (Petersen, 2001), although the variation in waste quality is large dependent on waste type and processing method (Smith et al., 1998). Optimal use of a waste product can benefit from an individual assessment of waste characteristics (Debosz et al., 2002). There are potential risks associated with recycling of urban waste materials. In addition to heavy metals, which are already regulated in most countries, it has become clear that organic contaminants like plasticizers and surfactants and their primary degradation products may accumulate, especially during anaerobic processing (Rogers, 1996). The potential for long-term accumulation of these contaminants in the soil, for plant uptake, and for disturbing soil functions needs to be addressed in more detail (O’Connor, 1996). Denmark has a long tradition for land application of sewage sludge, and until recently 70% was recycled. This figure has now decreased somewhat as a result of new intervention values for selected organic contaminants which in many cases prohibit recycling. In 1997, this development triggered a research program that was organized in the Centre for Sustainable Land Use and Management of Contaminants, Carbon and Nitrogen. The present report describes a joint field and greenhouse study in which several organic waste materials and contaminants were considered. Reference will also be made to related activities which have been carried out within the Centre for Sustainable Land Use and Management of Contaminants.

The investigation included (i) monitoring of soil and crop quality following a 3-year field trial with organic waste application; (ii) a 1-year plot experiment with banded application of sewage sludge that was designed to study the turnover and toxicity of sludge constituents under semi-realistic conditions; and (iii) a pot experiment with application of contaminants with or without sludge to study interactions among sludge, contaminants and soil. 2.1. Field sites Two field sites in Southwest Denmark were used for the 3-year field trial, a sandy loam near Askov Experimental Station (55◦ 28 N, 9◦ 07 E) and a loamy sand at the Lundgaard site located less than 5 km from Askov. Selected characteristics of the two soils are shown in Table 1. Climatic conditions were assumed to be almost identical. The 1960–1990 average annual temperature was 7.6 ◦ C, and average precipitation was 860 mm. 2.2. Waste materials and application rates Two types of sewage sludge, a household compost, and a solid pig manure were studied. Sewage sludge was obtained from a municipal wastewater Table 1 Selected soil characteristics for the two experimental sites Site

Soil type Clay (g kg−1 ) Silt (20–50 ␮m) (g kg−1 ) Sand (50–200 ␮m) (g kg−1 ) Sand (200–2000 ␮m) (g kg−1 ) C, total (%) N, total (%) P, total (mg kg−1 ) Pi (NaHCO3 ) (mg kg−1 ) K (NH4 acetate) (mg kg−1 ) Cu, total (mg kg−1 ) Zn, total (mg kg−1 ) pH(H2 O) (mg kg−1 ) CEC (meq 100 g−1 )

Askov

Lundgaard

Sandy loam 113 197 289 375 1.62 0.14 700 34 142 10 30 6.8 10

Loamy sand 52 81 211 632 1.43 0.1 560 24 93 5.8 19 6.3 6.7

S.O. Petersen et al. / Soil & Tillage Research 72 (2003) 139–152 Table 2 Selected characteristics of the waste products applieda SShigh (g kg−1 )

Dry matter 227 C (g kg−1 DM) 265 32.2 N (g kg−1 DM) NH4 + -N (g kg−1 DM) 5.5 31.5 P (g kg−1 DM) 1.8 K (g kg−1 DM) 106 Pbb (mg kg−1 DM) 2.2 Cdb (mg kg−1 DM) Hgb (mg kg−1 DM) 2.6 22 Nib (mg kg−1 DM) Crb (mg kg−1 DM) 30 977 Zn (mg kg−1 DM) 278 Cu (mg kg−1 DM) 9.15 PAHc,d (mg kg−1 DM) 55 DEHPc (mg kg−1 DM) 60 NPc (mg kg−1 DM) LASc (mg kg−1 DM) 2870

SSlow 162 276 57.8 11.2 33.0 4.9 68 1.4 0.9 18 18 447 215 3.38 27 12.5 110

Compost Manure 718 222 19.1 0.6 3.7 10.7 28 0.4 0.2 8 10 112 63 0.65 18 2.7 <10

237 337 31.2 10.6 23.7 17.9 – – – – – 1024 265 0.51 0.4 <0.5 <10

a The data represent mean value for waste applied in 1998, 1999 and 2000. b Data on Pb, Cd, Hg, Ni and Cr were derived from the continuous control at the waste treatment plants. c PAH: polycyclic aromatic hydrocarbons; DEHP: di(2-ethylhexyl)phthalate; NP:nonylphenol; LAS: linear alkylbenzene sulfonates. Current Danish intervention values for PAH, DEHP, NP + NPE and LAS are: 3, 50, 30 and 1300 mg kg−1 DM, respectively. d Naphthalene, acenaphtylene, acenaphthene, fluorene, phenanthrene, anthracene, fluoranthene, pyrene, benz(a)anthracene, chrysene, benzo(B + K)fluoranthene, benzo(a)pyrene, indeno(1,2,3cd)pyrene, dibenzo(a,h)anthracene, benzo(g,h,I)perylene.

treatment plant. Sewage sludge with a high load of organic contaminants (SShigh ; see Table 2) was derived from a pre-settling tank with chemical P removal; it had been anaerobically digested and dewatered prior to storage. Sludge with a low level of contaminants (SSlow ) was derived from an aeration tank and dewatered prior to use. Household compost (Compost) was obtained from a municipal composting facility processing kitchen waste; the waste was mixed with shredded straw (8 wt.%) to ensure adequate aeration during composting. Finally, solid pig manure (Manure) was obtained from a local pig operation. Table 2 presents information about nutrient composition and concentrations of selected contaminants. Organic matter content as well as nutrient composition was taken into account when choosing the waste application rates. For SShigh and SSlow , the annual

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rates corresponded to approximately 1000 kg C ha−1 . The nutrient availability of compost was limited, and solid manure typically has a high C:N ratio due to the bedding material included. In practical agriculture this would be compensated for by higher application rates and, accordingly, an annual rate of approximately 3300 kg C ha−1 was used for household compost and pig manure in the present study (Table 2). The actual application rates were based on analyses of dry matter content and volatile solids in six sub-samples taken upon receipt of the waste materials, and assuming 40% C in the organic matter. 2.3. Three-year field trial For this study, effects of the organic waste materials on crop yields and soil quality were evaluated in the final year of a 3-year crop rotation with spring barley (Hordeum vulgare, cv. Lamba), oat (Avena sativa, cv. Corrado) and spring barley (cv. Punto) in 1998, 1999 and 2000, respectively. No pesticides were used, and weeds were controlled mechanically as required. The four waste materials were applied either alone or in combination with supplementary mineral N (as calcium ammonium nitrate) corresponding to 80% of the expected crop N requirement. For reference, the experimental design included plots receiving mineral N alone at 40, 80 and 120% of the crop requirement, as well as an unamended control. All treatments were replicated three times in a randomized complete block design at both experimental sites. The overall plot size was 4 m × 14 m, within which a net plot for harvest of 1.5 m × 10 m was located. Wastes were distributed by hand and incorporated by rotovation. The experimental treatments were applied to the same plots every year. 2.3.1. Sampling of crop and soil Yield of the barley crop grown in 2000 was determined within the net plots for harvest, and plant samples were taken for analysis of contaminants from an area directly adjacent to the net plots. Soil samples from six selected treatments (the four waste materials without supplementary fertilizer, mineral N alone at 80% of crop requirement, and unamended soil) at both Askov and Lundgaard sites were analyzed for

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a range of physical, chemical and biological properties. At least 40 cores per plot (2 cm diameter, 20 cm depth) were taken and the soil pooled, the number of sub-samples being based of an analysis of variability using di(2-ethylhexyl)phthalate (DEHP) concentrations as a tracer for sewage sludge. DEHP was quantified in 10 samples taken within a selected plot (SShigh ) immediately after waste application. Since the distribution of DEHP was log-normal (0.05 < P < 0.10; Shapiro and Wilk, 1965), the variability test was carried out as described by Folorunso and Rolston (1982). Using 40 sub-samples, the DEHP (and thus sludge) concentration in the pooled sample was estimated to lie within 10–20% of the true mean with 95% confidence. 2.3.2. Soil micro-arthropods Intact soil cores (height, 5.5 cm; diameter, 6 cm) were sampled in triplicate from each of the treatments receiving SShigh , SSlow , compost, manure, mineral N or no amendments. Soil cores were extracted in a MacFadyen high temperature gradient extractor for 1 week. The collembolans in the extracts were classified into species according to Fjellberg (1980), while mites were classified into orders. 2.3.3. Soil fertility The fertility of waste-amended soil treatments was assessed by measurements of crop yields (average yields for the two sites are reported). Further, selected soil properties were quantified in the six treatments indicated in Section 2.3.1, including aggregate stability, extracellular polysaccharides, mineralizable N, plant available Pi , FDA hydrolysis activity, CO2 evolution and microbial biomass C. 2.3.4. Fate of contaminants Accumulation in the soil, as well as uptake of contaminants by the barley crop, was examined at the end of the 3-year crop rotation. Top soil was sampled at the start of the experiment in 1998, and each field plot was sampled after harvest in 2000. Soil samples from waste-amended and reference treatments were analyzed for polycyclic aromatic hydrocarbons (PAH; see footnote to Table 5 for details), DEHP, nonylphenol (NP)+nonylphenol mono- and diethoxylates (NPE), and linear alkylbenzene sulfonates (LAS), and inputs via organic wastes and manure were cal-

culated. For the crop, grain and straw were analyzed separately. 2.4. Banded sludge plot experiment In a separate experiment with a well-defined application of sewage sludge, the degradation of organic contaminants (i.e., loss of parent compound) in the sludge, and interactions between sludge and soil, were investigated under semi-realistic conditions. The experiment was conducted at the Lundgaard site, less than 200 m from the 3-year field trial. Two-dimensional sludge particles, i.e., sludge bands, that consisted of SShigh and SSlow were produced (length, 10 m; cross-section, 2 cm × 4 cm and 4 cm × 4 cm) and placed at approximately 6–10 cm soil depth. The spacing between bands was 0.5 m, and all treatments were replicated five times. The results shown represent at least three replicate samples. The two sludge materials, SShigh and SSlow , were applied in early May 1999, prior to sowing of oat, and sampling continued for up to a year depending on the analysis involved. Degradation patterns for 2 cm × 4 cm and 4 cm × 4 cm bands were similar, and only results for the latter are discussed. For sampling of the sludge-soil matrix, a rectangular corer was used which provided a 4 cm wide cross-section of the sludge bands and surrounding soil. Except where stated, the intact blocks were sectioned to give a sludge fraction and three soil fractions representing 0–1, 1–3 and 3–6 cm distance from the sludge at the depth of the sludge band. Sub-samples for chemical analyses were frozen until processed. The sludge fraction was analyzed for DEHP, NP + NPE and LAS in samples taken up to 1 year after sludge application. Effects of sludge constituents on the nitrification process were evaluated by measurements of potential ammonium oxidation (PAO) in the sludge and surrounding soil; only data for the 0–1 cm soil fractions and unamended soil are presented here. Turnover of the microbial community in/around sludge particles was investigated in bands with SShigh using phospholipid fatty acid (PLFA) analyses. The results shown here represent the center of the sludge band, as well as soil at 0.5–1.5 cm distance below and next to the sludge band. Sampling covered the 5-month period between application and harvest.

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2.5. Pot experiment A pot experiment was conducted with soil from Lundgaard. In one treatment, SShigh was added at a rate corresponding to 2000 kg C ha−1 , while in a second treatment (spiked) the contaminants (DEHP, NP and LAS) were introduced to the soil as an aqueous solution at concentrations equivalent to those of SShigh . The mixtures were homogenized using a blender. The pots were seeded with oilseed rape (Brassica napus, cv. Hyola), and the experiment was performed in triplicate. After 30 days growth, during which time the soil water content was maintained at approximately 60% of WHC, the rape plants were cut 1 cm above the soil surface for analysis of contaminant uptake. Soil in the pots was thoroughly mixed and sub-sampled for analysis of contaminants. 2.6. Analytical methods Soil structural stability was assessed as the fraction of soil in wet-stable aggregates >0.25 mm (Pojasok and Kay, 1990). Extracellular polysaccharides were measured as hot-water extractable carbohydrates (Haynes and Francis, 1993). Plant availability of P was estimated as resin extractable Pi (Rubaek and Sibbesen, 1993), henceforth resin-Pi . Nitrogen availability was quantified as N mineralized during a 7-day anaerobic incubation at 40 ◦ C (Keeney, 1982). The size and activity of the microbial biomass was estimated from fluorescein diacetate (FDA) hydrolysis

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activity (Schnürer and Rosswall, 1982), CO2 evolution rates were determined by gas chromatography, and biomass C was quantified, at Askov only, as described by Vance et al. (1987). Analytical methods primarily for analysis of plant material were developed for LAS, NP + NPE and DEHP (see Table 3). Mixtures of polar and nonpolar solvents were used for the extractions in order to break cell walls and dissolve the contaminants. Following clean-up, the extracts were concentrated to lower the detection limits. Soil samples were analyzed for DEHP, NPE and PAH using GC-MS selected ion monitoring (SIM), and for LAS using HPLC with fluorescence/UV detection (see Table 3). Quantification of DEHP, NPE and PAH was relative to surrogate standards biphenyl-d10 , pyrene-d10 , DEHP-d4 and benz(a)pyrene-d12 as appropriate. Quality assurance was by analysis of reagent blanks, of ‘spiked’ soil samples, and by inclusion of reference soil and sludge samples in each series. The detection limits were calculated from seven to eight replicate analyses of samples with concentrations at approximately five times the estimated detection limit, or from the variation of reagent blanks during internal quality control checks. PAO was measured in homogenized, sieved (<2.5 mm) soil samples using the chlorate inhibition technique (Belser and Mays, 1980) as modified by Hesselsøe et al. (2001a). Nitrite accumulation was linear (R2 > 0.9) during the first 8 h of incubation.

Table 3 Analytical methods adopted for analysis of (A) DEHP, NP + NPE and LAS in plant material and in soil from the pot experiment and (B) analytical methods for analysis of organic contaminants in waste materials and soil from the field experimentsa Component

Work up and concentration

Detection

(A) Plant samples/soil from pot experiment DEHP Hexane:acetone (1:1) NP + NPE Cyclohexane:acetone (9:1) LAS Water, methanol

Ultra Turrax, Florisil, evaporation Soxwave, NH2 , evaporation Sohxlet, C8, SAX, evaporation

GC-MS HPLC, fluorescence HPLC, fluorescence

(B) Waste materials/soil samples, field trial PAH Dichloromethane DEHP Dichloromethane NP + NPE Dichloromethane LAS pH > 12, methanol

Evaporation Evaporation Evaporation –

SIM GC-MS SIM GC-MS SIM GC-MS HPLC, UV and fluorescence

a

Extraction

Detection limit (C.V.) (mg kg−1 DM (%)) 0.1/0.05 (15) 0.1/0.01 (10) 0.2 (8) 0.001 (11) 0.05 (8) 0.05 (24) 10 (7)

PAH: polycyclic aromatic hydrocarbons; DEHP: di(2-ethylhexyl)phthalate; NP: nonylphenol; NPE: nonylphenol mono- and diethoxylates; LAS: linear alkylbenzene sulfonates.

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PLFA profiles were prepared as described by Nielsen and Petersen (2000) using phosphate buffer and chloroform for the extraction of lipids. A total of 31 fatty acids (C14 –C20 ) were identified. All results are reported on a dry weight basis. 2.7. Statistical analyses In the monitoring program conducted at the end of the 3-year field trial, treatments within each soil type were compared by one-way ANOVA followed by Tukey’s HSD multiple comparison test where appropriate. The mole percentage distribution of PLFA were analyzed by a principal component analysis (PCA) after log(x+1) transformation, and using the covariance matrix.

3. Results 3.1. Three-year field trial 3.1.1. Crop yields Fig. 1 shows total dry matter yield of barley in 2000 without and with supplementary mineral N corresponding to 80% of the crop N requirement (bars), as well as yield with increasing levels of mineral N fertilizer (curve). The organic waste rates applied gave

Fig. 1. Crop yields in soil amended with sewage sludge, household compost or solid pig manure without (plain bars) and with (hatched bars) supplementary mineral N corresponding to 80% of the crop requirement is shown. For reference, yields with increasing levels of mineral N alone were also obtained (curve). The results shown represent average yields for Askov and Lundgaard. Vertical bars indicate standard deviations.

satisfactory yields with SSlow alone, somewhat lower yields with SShigh and Manure, and a yield at the same level as unamended soil with Compost. Both sludge types and Manure gave satisfactory yields in the presence of supplementary mineral N, while the Compost treatment barely matched the corresponding yield with mineral N alone. 3.1.2. Effects on soil fertility indicators The proportion of soil in wet-stable aggregates (corrected for primary particles) was 3–9% at the coarse-textured Lundgaard site and 24–30% at the Askov site. There were no treatment effects (data not shown). At both sites, the concentrations of resin-Pi in the Manure treatment was significantly higher than in all other treatments (Fig. 2A), which reflected the higher P input level (78 kg P ha−1 per year as opposed to 20–40 kg ha−1 per year for the other treatments). At Lundgaard, the only other difference was a higher concentration in SSlow compared to Mineral N. At Askov, resin-Pi was significantly higher in SShigh and Compost than in the treatments with mineral N or no amendment, while SSlow was only elevated relative to the unamended soil. With respect to potentially mineralizable N (Fig. 2B) there were no significant differences at Lundgaard, while at the Askov site the level in the Compost treatment showed a significant accumulation of mineralizable N relative to SShigh and the unamended soil. FDA hydrolysis activity was stimulated in all waste treatments at Lundgaard (Fig. 2C), indicating that the organic inputs induced a sustained increase in microbial activity, even though the amounts of C introduced each year corresponded to only 0.07–0.2% of soil organic C. At Askov, SShigh and unamended soil had a significantly lower FDA hydrolysis activity than SSlow , Compost and Manure treatments. At Lundgaard, the Compost treatment had higher CO2 evolution rates than all other treatments except for Manure (Fig. 2D). At the Askov site, Compost had significantly higher respiration rate than the Mineral N treatment, while no other differences were significant. Extracellular polysaccharides showed no treatment effects at either site, concentrations at the two sites ranging from 317 to 417 mg C kg−1 . Biomass C at Askov ranged from 272 to 340 mg C kg−1 and also

S.O. Petersen et al. / Soil & Tillage Research 72 (2003) 139–152

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Fig. 2. Concentrations of resin extractable Pi (A), potentially mineralizable N (B), FDA hydrolysis activity (C), and CO2 evolution (D) at the end of the 3-year field trial in plots receiving organic wastes, mineral N or no amendments as indicated. Black bars represent Lundgaard; grey bars represent Askov. Statistically significant (P < 0.05) differences between treatments, but not across sites, are indicated by different letters.

showed no treatment effects at the time of sampling (data not shown). 3.1.3. Effects on soil fauna Relative to the unfertilized control, Compost and Manure gave rise to larger stimulations of micro-arthropod population densities than sludge and mineral N (Table 4). There were no discernable differences between the two sludge types. Stimulations in the coarse loamy sand at Lundgaard were considerably greater than in the sandy loam at Askov. 3.1.4. Persistence of contaminants in the soil Assuming no degradation of contaminants, soil inputs of DEHP, NPE, LAS and selected PAH compounds after three repeated applications of the organic wastes would be as shown in Table 5. The average yearly application rate was 3–4 Mg DM ha−1 for the two sludge types, 17 Mg DM ha−1 for household com-

post and 10 Mg DM ha−1 for pig manure. The highest addition of contaminants occurred with SShigh , except that the input of DEHP in Compost was of the same magnitude as in SShigh due to the higher application rate. The measured concentrations of LAS and NP+NPE in the soil were below the detection threshold in all treatments by the end of the 3-year field trial. Concentrations of DEHP ranged from below the detection threshold (<0.05 mg kg−1 ) to 0.103 mg kg−1 , and total PAH’s ranged from 0.125 to 0.209 mg kg−1 . There were no significant differences among treatments or between sites, except that the Askov soil with the higher clay and organic matter content had a higher background level of PAH’s. 3.1.5. Plant uptake of contaminants The barley crop grown in the last year of the field trial contained no LAS or NP in any treatment

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Table 4 Number of selected common species of collembola and mites by the end of the 3-year field trial (1000 specimens m−2 )a Control Askov Total micro-arthropods Collembola F. fimetaria I. notabilis I. palustris Mites Gamasida Lundgaard Total micro-arthropods Collembola F. fimetaria I. notabilis Mites Gamasida a

88.4 58.5 0.2 b 5.3 b 41.6 29.8 4.4 115.7 32.9 1.8 8.8 82.8 3.4

c c d b ab b

Mineral fertilizer

SShigh

SSlow

Compost

Manure

116.1 83.3 2.1 ab 12.6 ab 48.6 32.8 2.4

115.6 88.4 2.3 a 28.0 34.8 27.3 2.6

101.9 78.1 1.7 ab 12.9 ab 52.3 23.8 3.9

152.2 120.9 1.9 ab 46.5 a 46.6 31.3 3.1

120.5 86.5 3.1a 15.9 ab 43.5 34.0 5.0

144.5 71.0 2.4 15.9 73.4 5.0

231.2 143.4 20.3 66.4 87.8 7.8

181.0 101.2 7.3 50.9 79.8 8.1

295.9 187.8 7.4 114.9 108.1 5.0

333.8 216.3 39.9 67.7 117.4 9.6

c bc cd a b ab

ab a ab a ab ab

bc ab bcd a ab ab

ab a bc a ab ab

a a a a a a

Different letters within a row indicate significant difference (P < 0.05).

(Table 6). Similarly, no grain samples contained detectable concentrations of DEHP, while stem + leaves contained between 0.065 and 0.787 mg kg−1 DEHP, but with no relation to the amounts applied in the waste products (cf. Table 2). 3.2. Sludge band experiment The bands with SShigh were easily recognized in the soil throughout the 1-year period of the experiment. Crop roots started to colonize the sludge 3–4 weeks

after seeding, which may have stimulated aeration and possibly invasion by soil organisms. The degradation of DEHP in bands with SShigh was slow during the first 6 weeks of the experiment, and approximately 40% of the initial concentration was still present at the end of 6 months (Fig. 3). Degradation continued during winter, leaving only 5–6% of the initially applied concentration at the end of 12 months. In comparison with DEHP, the degradation of NP and LAS in SShigh were much faster with almost identical patterns of decline (Fig. 3). Around 70% of the initial

Table 5 Calculated inputs (mg kg−1 ) of DEHP, NPE, LAS and selected PAH compounds in the 3-year field trial after 3 years of repeated waste application, assuming no degradation had occurreda Treatment

SShigh

SSlow

Compost

Manure

Detection limit

PAHb,c

0.040 0.238 0.257 12.29

0.012 0.092 0.043 0.16

0.011 0.290 0.045 0.17

0.005 0.004 0.005 0.1

0.001 0.050–0.150d 0.050 10

DEHPb NPEb LASb

a The calculations were based on an even distribution at 0–20 cm depth, and a bulk density of 1.5 Mg m−3 . b PAH: polycyclic aromatic hydrocarbons; DEHP: di(2-ethylhexyl)phthalate; NPE: nonylphenol mono- and diethoxylates; LAS: linear alkylbenzene sulfonates. c Naphthalene, acenaphtylene, acenaphthene, fluorene, phenanthrene, anthracene, fluoranthene, pyrene, benz(a)anthracene, chrysene, benzo(B + K)fluoranthene, benzo(a)pyrene, indeno(1,2,3cd)pyrene, dibenzo(a,h)anthracene, benzo(g,h,I)perylene. d Two different analytical series had different detection limits.

Fig. 3. The disappearance of DEHP, NP and LAS was monitored in bands of SShigh during a period of up to 1 year. Vertical bars indicate standard deviations.

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Table 6 Concentrations (mg kg−1 ) of several organic contaminants in the barley crop grown in 2000, the last year of the 3-year field triala Compoundb

SShigh

SSlow

Compost

Manure

Mineral N

Askov DEHP Stem + leaves Grain NP + NPE LAS

0.108 <0.1 <0.1 <0.2

0.099 <0.1 <0.1 <0.2

0.109 <0.1 <0.1 <0.2

0.092 <0.1 <0.1 <0.2

0.099 <0.1 <0.1 <0.2

Lundgaard DEHP Stem + leaves Grain NP + NPE LAS

0.202 <0.1 <0.10 <0.2

0.787 <0.1 <0.10 <0.2

0.415 <0.1 <0.10 <0.2

0.065 <0.1 <0.10 <0.2

<0.1 <0.1 <0.10 <0.2

a b

Except for DEHP, the data represent both stem + leaves and grain. DEHP: di(2-ethylhexyl)phthalate; NP: nonylphenol; NPE: nonylphenol mono- and diethoxylates; LAS: linear alkylbenzene sulfonates.

amounts were degraded within the first 6 weeks, and <5% of NP and LAS remained after 6 months. Sludge bands with SShigh and SSlow both gave a 2–4-fold increase in PAO at 0–1 cm distance from the sludge compared to unamended soil, indicating an increased availability of ammonium (Fig. 4). Later in the growing season the stimulation was even larger, while PAO of the control soil remained constant. Hence, the stimulation was probably related to mineralization of organic N derived from the sludge, and not to climate or crop related factors. Overall changes in the distribution of PLFA in and around bands with SShigh were analyzed by PCA. Scores for the first three components explained

66.3, 9.3 and 6.5% of the total variation, respectively (Fig. 5A). The PCA was dominated by the difference in PLFA composition between sludge and soil. While there were no clear temporal PLFA changes in samples from the soil environment between May (white symbols) and October (black symbols), the PLFA composition of the sludge bands shifted towards that of the undisturbed soil in a gradual and nonlinear manner. The shift was partly caused by increasing proportions of cyclopropyl fatty acids, which are present in the cellular membrane of a range of primarily Gram negative bacteria, and which can be introduced in response to environmental conditions. Fig. 5B shows the ratio between cy17:0 and its metabolic precursor, 16:1␻7c, a ratio which has been shown to be inversely related to bacterial growth rate. This ratio in sludge was initially close to zero, but gradually increased to a value similar to that of the surrounding soil during the months following sludge application. 3.3. Pot experiment

Fig. 4. PAO at 0–1 cm distance from sludge bands with SSlow and SShigh , and in undisturbed soil. Vertical bars indicate standard deviations.

In the pot experiment, rape plants harvested 30 days after seeding contained neither DEHP, NP nor LAS at concentrations above the detection limits for the three contaminants (Table 3; see also Mortensen et al., 2001; Mortensen and Kure, 2003). The soil recovery of LAS, NP and DEHP in the SShigh and Spike treatments are presented in Fig. 6. The degradation of LAS and NP was delayed in sludge-amended soil compared to the Spike treatment, where contaminants were added in

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Fig. 5. (A) Principal component analysis based on PLFA profiles of SShigh (diamonds) and soil (squares) at the time of application in April (white), in October (black) and, for sludge, also at two times in June (light and dark grey, respectively). The PLFA composition of sludge gradually approached that of the soil. (B) Concentrations of cy17:0/16:1␻7c, a physiological stress indicator, in the soil (white) and in bands of SShigh (black). Vertical bars indicate standard deviations.

Fig. 6. Recovery after 30 days of LAS, NP and DEHP added to Lundgaard soil in SShigh or as an aqueous solution (‘Spike’). The pots were seeded with rape at the start of the pot experiment. Vertical bars indicate standard deviations.

pure form. In contrast, degradation of DEHP was faster when applied in a sludge matrix than when added directly to the soil. 4. Discussion Although some of the organic waste materials used in this study contained significant concentrations of several contaminants, all crop yields with supplementary mineral N (Fig. 1, hatched bars) were within satisfactory norms, except that some inhibition of growth

was suggested for Compost. No contaminant levels were elevated in this waste material, and partial immobilization of nutrients during spring is the most likely explanation for the lower yield with household compost. In the absence of supplementary fertilizer, the fertilizer value of SSlow was highest. The first year fertilizer value of organic wastes applied in spring is commonly ascribed to the amount of inorganic N, mainly as ammonium, in the waste material (Petersen, 2001). In the present study, SSlow and Manure treatments received similar inputs of inorganic N, but the yield was significantly higher with SSlow , presumably because mineralization of organic N contributed to the N availability in this treatment. Organic inputs may improve soil structure directly, by physical interactions with soil particles, and indirectly, by stimulating biological activity (Oades, 1993). Aggregate stability was not increased by the organic wastes in the 3-year field trial, although aggregation may have been transient (Debosz et al., 2002) or of a nature that is not sensitive to wet-sieving. León-González et al. (2000) found that compost application increased aggregation of a sandy soil as determined by dry-sieving, but for wet-stable aggregates there was no effect. Extracellular polymeric substances may help stabilize soil aggregates (Tisdall and Oades, 1982), but as for aggregation there were no treatment effects for extracellular polysaccharides 5 months after the final waste application.

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Resin-Pi had accumulated in manure-amended soil, while effects of sludge and compost amendment were not significant. Maguire et al. (2000) calculated that fertilization with biosolids to meet N requirements could lead to P saturation and environmental losses, although biosolids also increase the P sorption capacity of a soil. The recovery of resin-Pi from the waste materials in soil and crops constituted 40–50% of the input, and a large part of resin-Pi must therefore have become stabilized in less available forms, or leached from the top soil. Crop uptake of N from SShigh and Compost accounted for 33 and 13% of the input, respectively. The retention of compost N in organic forms was reflected in the accumulation of potentially mineralizable N at the Askov site, but not at Lundgaard (Fig. 2B). In an associated incubation experiment it was found that net N mineralization during the first year was 6-fold lower from Compost than from SShigh , even though the N input with Compost was 2.5 times higher (Debosz et al., 2002). The greatest increases in soil micro-arthropod population densities were seen with Compost and Manure, which also represented the highest C inputs. It is conceivable that microbial growth derived from the waste material increased the food availability for soil fauna. The stimulation was far greater at Lundgaard than at Askov, which we believe could be related to the different soil textures at the two sites (cf. Table 1). The protective pore space, i.e., pores with access to bacteria but not to grazers (Postma and van Veen, 1990), was estimated to be 50% larger at Askov than at Lundgaard (data not shown), and also the fraction of large pores defined by the sand fraction would influence the ability of collembola and mites to explore the food source. The increase in potentially mineralizable N observed at Askov, but not at Lundgaard, also indicated a better protection of labile organic matter at the Askov site. Organic waste amendments stimulated the microbial activity, as reflected in FDA hydrolysis activity (a broad index of hydrolytic activity) and, to some extent, CO2 evolution rates (cf. Fig. 2C and D). This was consistent with strongly increased rates of basal and substrate induced respiration around sludge bands (K. Brandt et al., unpublished results). Effects on soil microbial functions were studied in the sludge band experiment, as represented by ammonium oxidation, which is carried out by a relatively

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uniform group of bacteria considered to be sensitive to many pollutants, including LAS (Brandt et al., 2001). Although SShigh and SSlow contained very different amounts of contaminants, the two sludge types stimulated soil PAO immediately outside the sludge particles to the same degree (cf. Fig. 4). This indicates that contaminants in the sludge had no adverse effects on the nitrification process, although it must be stressed that the PAO rates recorded represented net values that could be composed of simultaneous stimulatory and inhibitory effects (Elsgaard et al., 2001b). Concentrations of LAS in SShigh and SSlow were 2870 and 110 mg kg−1 DM, respectively (Table 2). An EC10 value for PAO of <8 mg LAS kg−1 soil was determined by Elsgaard et al. (2001a), which implies that, under the experimental conditions employed, LAS was effectively retained (and degraded) within the sludge material. This conclusion is consistent with data from a parallel experiment with SSlow initially containing 31,000 mg LAS kg−1 , which used the same experimental design and study area as presented here; only very short-lived (1–2 weeks) toxic effects of LAS on PAO were observed in soil sampled 0–1 cm from the sludge (Brandt et al., 2000). The continued increase in PAO during the summer (Fig. 4) suggested that there was a considerable mineralization of sludge-derived N. The changes in PLFA composition of the sludge material (Fig. 5A) indicated that microbial succession (possibly promoted by colonization via roots, or by fractures induced by drying-rewetting or bioturbation) altered the microbial community in the direction of a community more similar to that of the surrounding soil. The low initial cyclopropyl-to-precursor ratios in the sludge fraction (Fig. 5B) suggested that SShigh was dominated by actively growing organisms (Grogan and Cronan, 1997). The subsequent increase during the following months would then suggest that SShigh was gradually depleted of easily degradable substrates, leading to cyclopropyl-to-precursor ratios more similar to those of the nutrient-poor soil environment. This interpretation is in accordance with respiration measurements in the sludge, which showed decreasing microbial activity during the growing season (K. Brandt et al., unpublished results). The environmental fate of contaminants in organic wastes applied to soil, including the potential for plant uptake, is a matter of concern. However, neither

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sludge, household compost nor manure applications resulted in statistically significant accumulations of any of the organic contaminants analyzed (data not shown). This is consistent with other field studies using sludge applications at moderate rates, in which no accumulation of LAS, NP or DEHP was observed (Mieure et al., 1990; Vikelsoe et al., 1999). The calculated inputs should, at least with SShigh , be detectable, which indicates that contaminants were likely degraded for the most part. Previous laboratory studies have indicated that sewage sludge contains a microflora with a potential to degrade NP, DEHP and other organic contaminants, and that the actual degradation occurs within the sludge when distributed as discrete particles (Roslev et al., 1998; Madsen et al., 1999; Hesselsøe et al., 2001b). However, LAS and NP are not ultimately degraded and DEHP degradation is strongly reduced under anaerobic conditions (Gejlsbjerg et al., 2001), suggesting that the aeration status of the sludge material is a key factor in determining the time course of contaminant degradation. Oxygen penetration into the sludge band was examined with micro-electrodes in a related study (Hesselsøe et al., 2001b). Oxygen penetration at the end of 6 weeks averaged 2 mm, while at the end of 6 months it varied between 4 mm and full penetration. Hence, during the first several weeks contaminant removal within the sludge was probably limited to a 2 mm zone at the sludge–soil interface. At the end of 6 weeks, approximately 15 and 30% of the initial LAS and NP, respectively, were recovered in the sludge bands (see Fig. 3), indicating that a considerable part of the degradation occurred after diffusional transport from anaerobic to aerobic sites. The stronger adsorption of DEHP (log Kow = 7.6; see below) would prevent diffusional transport and reduce bioavailability, which may explain the longer turnover times for this contaminant. In the 3-year field trial, the size of sludge particles incorporated was typically 2–4 cm in diameter, and the controlled application of sludge as 4 cm × 4 cm bands therefore represented a worst case scenario with respect to oxygen availability. Still, residual concentrations at the end of the growing season in the sludge band experiment could be extrapolated to the 3-year field trial with some confidence. Residual concentrations of LAS and NP after 6 months were below 5% of the initial levels, and DEHP below 40% of the ini-

tial level. After 1 year, residual DEHP was less than 10% of the concentration in the sludge applied. Even if no further degradation were assumed, the resulting increase in LAS, NP and DEHP during 3 years would be below the detection threshold. The affinity for binding of contaminants to soil organic matter is important when assessing contaminant mobility and its potential for plant uptake. This property is generally described by the partition coefficient octanol-to-water (Kow ). Duarte-Davidson and Jones (1996) found that, for compounds with log Kow > 4, the potential for root retention was high and the potential for uptake and translocation low. LAS was the most water-soluble contaminant investigated, with a log Kow of ca. 3.7. In earlier experiments where sludge with radio-labeled LAS was applied to soil, about 6% of the label was recovered in plant biomass, but no distinction was made between the parent compound and metabolites (Figge and Schöberl, 1989). In the present study a specific analytical method was used, and no LAS could be detected in the green parts of rape in the pot experiment, nor in the final barley crop of the 3-year field trial. Only in a separate experiment, where sludge was spiked with LAS to give a resulting initial concentration of 31,000 mg kg−1 , was LAS detected in oat leaves at a concentration of 2.2 mg kg−1 (data not shown). log Kow of NP and DEHP were 6.2 and 7.6, respectively. There was no plant uptake of NP; all analyses of plant material were below the detection limit in the field trial, as well as in the pot experiment. DEHP was detected in the green parts of barley at the end of the field trial, although this compound was the least polar contaminant investigated. Previous studies have not seen DEHP uptake from sludge-amended soil (Aranda et al., 1989; Grøn et al., 2001) and there was no relationship between waste and crop concentrations of DEHP, Hence, it was concluded that the plant content of DEHP was mainly derived from atmospheric deposition (Grøn et al., 2001).

5. Conclusions This study represented a multi-facetted evaluation of the short-term effects of organic waste application to agricultural soil. There was considerable difference in the fertilizer value of the waste materials, and one

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of the sludge types (SSlow ) had a fertilizer value exceeding that of solid pig manure. The organic wastes generally showed no adverse effects on crop yields, soil fertility or biological activity, but rather a stimulation of some properties. Ammonium oxidation was stimulated within a few mm distance from sludge with toxic levels of LAS, which suggested that contaminant degradation in field-applied organic waste was to a large extent closely associated with the waste. The levels of typical organic contaminants decreased considerably within a period of 6–12 months after the time of application, and the monitoring carried out at the end of the 3-year field trial gave no indications of long-term accumulation. Also, there was no evidence for plant uptake of the organic contaminants studied. In conclusion, under current Danish regulations the risks associated with recycling of municipal organic wastes to agriculture appear to be limited, and potential uses for individual waste products should be further explored.

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