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GEODER-11950; No of Pages 11 Geoderma xxx (2015) xxx–xxx

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Land use and soil factors affecting accumulation of phosphorus species in temperate soils Marc I. Stutter a,b,⁎, Charles A. Shand a,b, Timothy S. George a,b, Martin S.A. Blackwell c, Liz Dixon c, Roland Bol c,h, Regina L. MacKay d, Alan E. Richardson e, Leo M. Condron f, Philip M. Haygarth g a

The James Hutton Institute, Aberdeen AB15 8QH, UK The James Hutton Institute, Dundee DD2 5DA, UK Rothamsted Research North Wyke, Okehampton, Devon EX20 2SB, UK d College of Life Sciences, University of Dundee, Dundee DD1 5EH, UK e CSIRO Plant Industry, Black Mountain, Canberra ACT 2601, Australia f Faculty of Agriculture and Life Sciences, Lincoln University, Lincoln 7647, Canterbury, New Zealand g Lancaster Environment Centre, Lancaster University, LA1 4YQ, UK h Institute of Bio and Geosciences, Agrosphere, Forschungszentrum Jülich IBG-3, 52428 Jülich, Germany b c

a r t i c l e

i n f o

Article history: Received 10 July 2014 Received in revised form 30 January 2015 Accepted 17 March 2015 Available online xxxx Keywords: Phosphorus species Soils Carbon Oxalate extractable Fe, Al Land use

a b s t r a c t Data on the distribution of phosphorus (P) species in soils with differing land uses and properties are essential to understanding environmental P availability and how fertiliser inputs, cropping and grazing affect accumulation of soil inorganic P (Pi) and organic P (Po) forms. We examined thirty-two temperate soils (with soil organic C concentrations 12–449 g C kg−1 and total P 295–3435 mg P kg−1) for biogeochemical properties of soil C, reactive surfaces and P by common indices and 31P-NMR spectroscopy on NaOH–EDTA extracts for P species. Arable soil P was dominated by inorganic orthophosphate (276–2520 mg P kg−1), N monoester P (105– 446 mg P kg−1). The limited diesters, polyphosphates and microbial P in arable soils suggest that cropping and fertiliser inputs limit ecosystem microbial functions and P diversity. Intensive grassland had inorganic orthophosphate concentrations (233–842 mg P kg− 1) similar to monoesters (200–658 mg P kg− 1) N diesters (0–50 mg P kg−1) and polyphosphates (1–78 mg P kg−1). As grazing became more extensive P in semi-natural systems was dominated by organic P, including monoesters (37–621 mg P kg−1) and other diverse forms; principally diester (0–102 mg P kg − 1 ) and polyphosphates (0–108 mg P kg − 1). These were related to SOC, water extractable organic carbon (WEOC) and microbial P, suggesting strong microbially-mediated processes. A number of abiotic and biotic related processes appeared to control accumulation of different soil P species and gave considerable variability in forms and concentrations within land use groups. The implications are that to increase agricultural P efficiencies mechanisms to utilise both soil Pi and Po are needed and that specific management strategies may be required for site-specific circumstances of soil C and reactive properties such as Fe and Al complexes. © 2015 Elsevier B.V. All rights reserved.

1. Introduction Phosphorus (P) is a critical, non-renewable resource crucial to agricultural productivity (Neset et al., 2008; Stutter et al., 2012). Concerns

Abbreviations: SOC, soil organic carbon; WEOC, water extractable organic carbon; Pi, inorganic P; Po, organic P; WEPi, water extractable inorganic P; WEPo, water extractable organic P; Mox, element ion extracted by acid ammonium oxalate; Psat, the molar ratio Pox/ (Feox + Alox); Po-citric, organic P extracted by citric acid; P-NaOH–EDTA, total P extracted by NaOH–EDTA; 31P NMR, soil P compounds; MDP, methylene diphosphonic acid; Ortho Pi, inorganic orthophosphate; P-mono, orthophosphate monoesters; P-IHP, inositol hexaphosphate monoester P; P-di, orthophosphate diesters; P-polyp-end and P-polyp-mid, end and middle group polyphosphates, respectively; P-phosphon, phosphonates. ⁎ Corresponding author. E-mail address: [email protected] (M.I. Stutter).

for food security (Godfray et al., 2010) necessitate an improvement in P efficiencies of agronomic systems based on specific knowledge of the nature of P species in soils, turnover rates and bioavailability. Agronomic practices exert major influences on the natural P balance (McDowell and Stewart, 2006). Over the past century there has been a build-up of P in the soils of many regions of the world by the addition of fertilisers in amounts exceeding offtake (Neset et al., 2008; Sattari et al., 2012). Chemical fertilisers derived from rock phosphate have played a key role in the agronomic revolution. Although the size of global mineral P resources is disputed, it is a finite resource and requires considerate utilisation. An improved knowledge of the abundance and diversity of soil P species and of the soil and land use factors that have governed the pathway to current soil P conditions will inform better management to benefit agricultural production, water quality and soil quality. 0016-7061/© 2015 Elsevier B.V. All rights reserved.

Please cite this article as: Stutter, M.I., et al., Land use and soil factors affecting accumulation of phosphorus species in temperate soils, Geoderma (2015),


M.I. Stutter et al. / Geoderma xxx (2015) xxx–xxx

In many soils, P added as fertiliser becomes ‘fixed’ by sorption to mineral phases and converted to compounds of more limited bioavailability. Thus, continual application of fertiliser P is required to ensure that a small proportion remains free to sustain yields and research effort is now being directed at improving the acquisition of soil P by crop plants (Fageria et al., 2008; Ramaekers et al., 2010; Richardson et al., 2011). Increasingly it is being recognised that organically-complexed P (Po) compounds are an important component in the maintenance of P supply to crops from soil, but additional research is required in order to fully understand and utilise these compounds (Stutter et al., 2012). There is a growing body of literature on the characterisation of Po species in agricultural soils especially by NMR. Yet, there is a need for better understanding of the importance of Po across a range of agricultural and natural systems with a continuum of soil P inputs and carbon (C) concentrations. The cycling of P and C in soils is likely to be tightly coupled since bioavailability of C drives ecosystem processes in turn affecting sequestration and supply of soil P, with associated change in soil P species (Gressel et al., 1996; Bradford et al., 2008; Spohn and Kuzyakov, 2013; Kirkby et al., 2013). This P:C macronutrient coupling is especially important to understand with current efforts to increase soil organic C storage for benefits to soil biodiversity, aggregate stability and climate policy. The accumulation of different soil P compounds is controlled by soil–plant-microbial P cycling, in turn dependent on environmental and management conditions. In natural systems, a balance is attained by interaction between mineral weathering, sorption/desorption of P compounds bound to soil surfaces, and the mineralization–immobilization P balance (Dalal, 1977; Condron et al., 2005). In the context of our study we define mineralization and immobilization as the replenishment and draw down, respectively, of cropavailable inorganic P forms in soil solution and potentially-soluble soil complexes. The microbially-mediated process of P immobilization is generally thought to lead, alongside abiotic stabilisation processes, to accumulation of Po forms (Bünemann et al., 2004) but there is also potential for immobilized Pi associated with SOC (Turner et al., 2006). Within both the inorganic and Po fractions in soils there are gradients of bioavailability and transport mobility (Frossard et al., 2000). Organic-P can be the dominant fraction in temperate soils and exists in several different chemical species. It is important to have a good knowledge of Po compounds in soils because a large proportions of inorganic P fertilisers added to soil can be transformed to Po (Condron et al., 2005). Nuclear magnetic resonance spectroscopy (NMR) allows quantification of soil P species (Turner et al., 2003a,b; McDowell et al., 2006; Doolette et al., 2011; Doolette and Smernik, 2011) and coupled with analyses of wider soil properties has potential to open up the new insight into soil P turnover processes necessary to better manage the soil P resource. Sundareshwar et al. (2009) recently reported on the use of NMR techniques to determine the ‘soil P diversity’, positively relating the occurrence of a wider range of soil P compounds with enhanced soil biogeochemical function and ecosystem services (such as nutrient retention and soil C stability) across land use gradients. Our hypothesis was that soil P accumulation, including the concentrations and proportions of different P species, will show strong variation with land use, being primarily influenced by land management factors. To address this we used 31P NMR to study changing Pi and Po compositions and compared these data to commonly used P indices for plant availability and environmental mobility. Furthermore we proposed that differences in soil parameters (soil organic C and surface reactivity properties both within and between land use classes) would additionally explain variation in soil P accumulation that comprised microbially-influenced and abiotic stabilisation processes. The study examined thirty-two temperate topsoils from dominant UK land use classes and sought to maximise variation in combinations of C and P inputs.

2. Materials and methods 2.1. Soil sampling Topsoils were collected in spring 2008 from 32 sites in the UK representing the dominant agronomic land uses designed to capture a wide range of organic C and P contents (Supplementary Information Tables S1 and S2). Sites included arable (mainly winter wheat, some barley and oats), intensive grassland (a mixture of lowland intensive pasture and two fertilised and grazed upland sites), and a mixed group categorized as ‘extensive grazing/semi-natural’ (comprising lightly grazed lowland rough grassland, marsh to upland moorland, but including a single forested site). Additionally arable buffer sites were rough grassland, typically set-aside or buffer strip areas adjacent to arable sites which had not recently been cultivated. Necessarily this provided an unbalanced sample design with respect to C and P across land use classes as Histosols were present only under moorland (low P input) but not arable (high P input) land use. Grid-sampling was undertaken at 2-m spacing across an 8 by 8 m compass-aligned grid and a composite sample made of the 25 spatial subsamples. The soil samples were taken from 0–15 cm depth for arable and arable buffer sites, and 0–7 cm depth for grassland and extensive grazing sites (including the forest). The different sampling depths for arable (0–15 cm) compared to grassland and extensive soils (0–7 cm) followed the standard methodology for agronomic testing of UK soils (Defra, 2010). The rationale for this is that for arable soils (and previously for arable buffer soils) ploughing homogenises soils over this depth but stratified pasture and semi-natural soils should be sampled at shallower depths to better inform runoff and plant P availability studies (Vadas et al., 2005). Soil samples were sieved (b2 mm) whilst field moist, subsampled for the microbial P analysis and the remainder air-dried (30 °C) until stable mass was attained on successive days (between 3 and 7 drying days). Air-drying is common practice in soil chemical and 31P NMR analysis preparation (Cade-Menun and Liu, 2014). These authors' review noted no consistent reported effects of different drying regimes on P speciation and concentrations and recommended consistency in approach. For analysis of CN and oxalate Fe, Al subsamples were prepared by grinding the air dry soil in an agate ball mill to a particle size b 150 μm. 2.2. Physico-chemical analyses of soil Total N and organic C concentrations were measured using a CHN analyser (model Flash EA 1112 CN, Thermo-Finnigan, Italy). Soil pH was measured in 0.01 M CaCl2 using a 1:3 soil to solution ratio. Oxalate extractable elements (Pox, Alox, Feox) were determined by extraction of the soil with acid ammonium oxalate (Farmer et al., 1983) and analysis of the extract by inductively coupled plasma optical emission spectroscopy (ICP-OES, Agilent 7500ce instrument, Agilent Technologies, CA, USA). The P saturation index was calculated as Psat = [Pox] / ([Alox] + [Feox]), where [] denotes the element's molar concentration. Total P was determined by a NaOH fusion method (Smith and Bain, 1982) comprising ignition of 0.1 g samples with solid NaOH, titrating the melt to pH 7 with 20% H2SO4 and subsequent P analysis by ICPOES. Olsen P was determined according to Sim (2000). Microbial P was measured by a procedure modified from McLaughlin and Alston (1986). Replicate 2 g samples of field-moist soil were extracted for 16 h in 20 ml of deionised water with anion exchange resin strips either with or without addition of 0.8 ml hexanol. The phosphate collected in resin strips was eluted with 0.1 M HCl and the concentrations of molybdate reactive P determined colorimetrically. Microbial P (Microb-P; detection limit b0.5 mg P kg−1) was estimated as the difference between samples extracted with and without hexanol. A correction factor to account for sorption of P during extraction was determined from samples spiked with 20 mg P g−1. Colorimetric determinations for Pi used the malachite green complex (Irving and McLaughlin, 1990).

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Water extractable P and water extractable organic C (WEOC) were determined by shaking the 2-mm sieved soil with deionised water (1:4 w/v) for 1 h. After filtration (Whatman 42) the P in the extract was subdivided into inorganic (WEPi) and organic (WEPo) P colorimetrically and DOC was determined (TOCV-CSH, Shimadzu, Japan). The Fe oxide paper strip method (Chardon et al., 1996) with subsequent colorimetric analysis for Pi was used to determine P desorbable to a weak Psink (termed Fe–P). Citric acid extractable organic P (Po-citric) was determined by extraction of the 2-mm sieved soil at 1:2 w/v with 50 mM citric acid at 22 °C for 30 min on a reciprocal shaker (300 rpm). The extracts were centrifuged (10,300 g) for 15 min and the Pi concentration determined colorimetrically (Skalar San++, the Netherlands) and a separation of Pi and Po made colorimetrically using the malachite green complex (Irving and McLaughlin, 1990). For this the total P concentration was determined by autoclave persulphate digestion (Schoenau and Huang, 1991), and Po in the extracts calculated as total P − Pi Full soil data are presented in Supplementary Information Table S2. 2.3. 31P NMR analyses The 2-mm sieved soil was extracted on an end-over-end shaker (20 °C, 16 h) with a solution of 0.25 M NaOH and 0.05 M EDTA at a 1:20 w/v ratio (Turner et al., 2003b). Recent work (Cade-Menun and Liu, 2014) suggests that smaller extract ratios of 1:10 or 1:5 improve P recoveries for mineral soils but we applied 1:20 across our range of soils. The extract was subsampled for the determination of the total P, Fe and Mn concentration by ICP, and the remainder of the extract freeze-dried. The recovery of P by NaOH–EDTA soil extracts (Supplementary Information Table S3) was compared to total P (after the NaOH fusion method) and showed mean (±1 S.E.) recovery 72 ± 4%, with no significant difference between the four land use classes (p = 0.48). Extraction efficiencies (Table S3) b 100% are expected to represent acid soluble P predominantly in a mineral matrix (Turner et al., 2007; referred to here as ‘unconfirmed’ P species in Table S3). Two values N 100% demonstrate slight analytical errors. For NMR analyses and immediately prior to placing the sample in the spectrometer, 100 mg of the freeze-dried material was dissolved in 1 ml of 1 M NaOH containing 10% v:v D2O for frequency locking. The solution was centrifuged (to remove particles that would contribute to line broadening). Spectra were acquired using an Avance 500 II instrument (Bruker, Germany) operating with proton decoupling at 203 MHz with a 5 mm probe. We used a 90° pulse angle with acquisition time 0.9 s, and pulse-delay 1–2 s operating at 21 °C. Delay times of 1–2 s at 90° pulse are small for accurate quantification of P species, potentially negatively biasing diesters and positively biasing orthophosphate (McDowell et al., 2006). For a subset of four soils (differing in P/ Fe + Mn ratios) we found little evidence of impact of changes in relaxation times between 1 and 10 s (Supplementary Information Table S4 and Fig. S3). We were also able to determine T1 values for the ortho-Pi peak using inversion recovery experiments and calculations of the null point (Keeler, 2010). For Vealand improved, Southlake South and Tayport T1 was 0.72, 0.14 and 0.87 s, respectively indicating recovery times ~2 s to achieve equilibrium in 95% of spin populations for those tested (Keeler, 2010). Times for T1 could not be determined for the Alness 2 soil due to a weak signal. The P/(Fe + Mn) mass ratios (Supplementary Information Table S3) with a mean ± 1 S.E. of 3.1 ± 0.9 significantly differed between the land use classes (ANOVA; p b 0.001). Arable soils had higher P/(Fe + Mn) ratios necessitating relaxation times of 6.1 ± 1.8 and 3.1 ± 0.4 to achieve 3 × T1 times for Pi and Po quantification, respectively, according to the regression relationships of McDowell et al. (2006). Shorter mean 3 × T1 times were indicated necessary for the other land use classes (1.5 to 2.2 s for Pi and 2.2 to 2.3 s for Po). The numbers of scans was fixed at 10,200 to standardise experimental times with respect to potential degradation times of certain


susceptible components of P-di in the alkaline extract (Turner et al., 2003a,b). Sample peak positions were compared to that for an internal standard comprising a fixed mass of methylene diphosphonic acid (MDP). The MDP was introduced with each sample within a sealed glass capillary tube to guard against possible degradation in the sample matrix (Bedrock et al., 1994). This provided a relative standard position on the ppm scale. Additionally the sum of peak areas of all P forms when normalised to the MDP peak for each sample gave a highly significant relationship with total P in the NaOH–EDTA extracts (see Results). Yet a check on quantification using the MDP peak gave a consistent bias, which we believe was due to the way the MDP was introduced with each sample with part of the capillary not within the instrument's NMR field (ie partly protruding above the sample). We set chemical shifts relative to the peak for the MDP standard at 16.8 ppm and used conventional chemical shift assignments from the literature (Turner et al., 2003a, 2012; McDowell et al., 2006; Giles et al., 2011; Vincent et al., 2013): Inorganic orthophosphate (Pi) 6.06 to 6.22; phosphate monoester group (P-mono) 4.10 to 5.97; phosphate diesters (P-di) − 0.2 to − 0.3; phosphonates (P-phosphon) 16.54 and 20.60; polyphosphate end groups (P-polyp-end) − 4.23 to − 4.37. These assignments became the main groups that were used in statistical analyses. In addition several other assignments were made. Whilst data for these species concentrations appear in Supplementary information Table S3, they were either not used for statistical purposes or grouped: polyphosphate middle groups (P-polyp-mid) −20.51 were identified only in the single Dartmoor sample these were not considered in statistical analyses; shifts at 6.64 to 6.78 were taken as phytate stereoisomers according to Turner et al. (2007) and were added to the P-mono total (comprising b2% of total P in any soil); the specific shift at 4.00 was attributed to scyllo forms of P-IHP and were added into the P-mono group. By spiking the extracts of samples 5, 18 and 32 with myo-inositol hexaphosphate dodecasodium salt solution to approximately double the concentration we identified the presence of inositol hexaphosphate (P-IHP; a specific form of P-mono) as shifts at 4.48, 4.63, 4.99 and 5.88 ppm. However we did not separately quantify P-IHP in this study nor account for the degradation of P-di by reapportioning some of the monoester shifts back into the quantified diester pool. Other authors have shown across the NMR literature (see review by Cade-Menun and Liu, 2014 and references therein) that certain phosphate diesters are prone to degradation in alkaline extracts. Although we are aware that peaks for β-glycerophosphate (~4.9 ppm) are currently understood to be degraded phospholipids (Doolette et al., 2011) and generally wrongly assigned to the P-mono group we did not correct for these minor peaks. We did not separate shifts for pyrophosphate from those of the general polyphosphate end group (P-polyp). When normalised to the internal MDP standard (16.8 ppm) the sum of peak areas identified for the soil P compounds was significantly related to total P in the extract (P-NaOH–EDTA; mg P kg−1) measured by inductively coupled plasma-optical emission spectroscopy (ICP-OES), (relative peak area = 0.0066 ± 0.0002 ∗ P-NaOH–EDTA − 0.017 ± 0.23; R2 = 0.97; p = b 0.001; n = 32, with 1 S.E. error terms). For reasons explained in the methods were unable to use the MDP mass introduced with each sample as an internal quantification reference. Example spectra are given in Fig. 1 for (a) the Woburn arable, (b) Canol intensive grassland and (c) Dartmoor extensive grazing, moorland soils. The full set of spectra is presented in Supplementary Information Fig. S1. NMR P species concentrations were expressed as percentages of the total soil P by NaOH fusion and add to 100% with the inclusion of the ‘unextracted, alkaline insoluble P’ fraction (Table S3). 2.4. Data analysis and statistical methods Processing of NMR spectra was performed with MestReNova v8.1. Statistical analysis was performed using Minitab v16. Where necessary, according to results of Andersen–Darling test for normality, data were

Please cite this article as: Stutter, M.I., et al., Land use and soil factors affecting accumulation of phosphorus species in temperate soils, Geoderma (2015),


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(a) Woburn 2, arable



7 6 5 4 ppm E

(b) Canol, intensive grassland



G 7 6

5 4 ppm


(c) Dartmoor, extensive


7 6 5 4 ppm I

22 18 14 10 6


-2 -6 -10 -14 -18 - 22 ppm

Fig. 1. Example whole scale 31P NMR spectra with inset expanded monoester regions for (a) Woburn 2 arable, (b) Canol intensive lowland pasture and (c) Dartmoor extensive moorland soils. The shifts identified are: A, phosphonates (P-phosphon); B, MDPA standard; C, undefined phytate stereoisomers; D, inorganic orthophosphate (Pi); E, phosphate monoester group (P-mono); F, scyllo phytate forms; G, phosphate diesters (P-di); H, polyphosphate end groups (P-polyp-end); I, polyphosphate middle groups (P-polyp-mid).

log10 transformed (n, or n + 1) prior to ANOVA and correlation analyses. Given the unequal group sizes for land use classes Levene's test was used to confirm that the hypothesis of inequality of variances could be rejected (p N 0.05). The Tukey test was used following a General Linear Model ANOVA to identify significant differences between land use classes (p b 0.05) that are denoted by different superscript letters between groups (Table 1). Principal components analysis, using the correlation matrix, was used to derive a non-biased reduction in the analytical parameters of NMR P species, P indices and other soil properties, show the separation in the PC1 vs PC2 space according to land use classes and the clustering of the loading parameters against the soil PCs using a biplot. 3. Results 3.1. Variability in soil properties and P indices Soil properties, P indices and P species determined by 31P NMR are shown for individual sites (Supplementary Information Tables S2 and S3, respectively) and by land use classes (Table 1). Soil pH ranged from 3.28 (Dartmoor, extensive) to 7.1 (Woburn Broadmead, arable field). Soil organic carbon contents ranged from 12 g C kg−1 (Woburn stackyard, arable) to 449 g C kg−1 (Dartmoor, extensive grazing). Soil organic carbon concentrations significantly differed across land use class (p b 0.001; Table 1) in the order: extensive land use (158 ± 130 g C kg−1; mean ± 95% confidence interval) N intensive grassland (51 ± 15 g C kg−1) N arable (22 ± 4 g C kg−1). The considerable variation in SOC for the extensive land use class occurred since this category contained low P input extensive grazing Gleysols with low to medium

SOC, the forested low SOC Podzol and two moorland Histosols. Organic C:total N mass ratios were generally low (mean 9.8) but highest for the Dartmoor soil (17.3) and did not differ significantly between land use class (p = 0.31). Olsen P ranged from 6 mg P kg−1 (Tentsmuir, the only forested site) to 184 mg P kg−1 (Haddington 1, arable), being significantly different between land uses (ANOVA p = 0.001). In the UK, the Olsen P measurement provides a regulatory index for fertiliser recommendations (DEFRA, 2010). Using this index 95% of soils under arable, 75% under intensive grassland, and 25% of soils under extensive land use would be advised to receive fertiliser P at rates not exceeding crop offtake. Iron oxide paper strip P (Fe–P) was significantly positively related to Olsen-P (R2 = 0.56; p b 0.001) but did not significantly differ across land uses (p = 0.19). Molar Psat (denoting saturation of Alox and Feox soil exchange complex with P) varied significantly with land use (ANOVA p = 0.001) with significantly greater values for arable than other land use classes, but Alox and Feox concentrations did not differ significantly with land use (p = 0.89 and 0.11, respectively). There was a highly significant relationship between Olsen P and Psat (Olsen P = 429 ∗ Psat + 4.9; R2 = 0.84, p b 0.001). Citric acid extractable organic P (Po-citric; dissolved organic P desorbed from the soil by low molecular weight organic acid) did not differ significantly with land use (p = 0.58). Microbial biomass P (Microb P) was significantly related to log10SOC (R2 = 0.44, p b 0.001) and differed significantly between land use classes (ANOVA p = 0.03). There were large values of Microb P for extensive soils from Dartmoor and Southlakes South (63 and 19 mg P kg−1), for intensive grassland soils from Banadl and Southlakes North (23 and 17 mg P kg− 1), but also for the arable Rothamsted soil (12 mg P kg−1). Water extractable P was subdivided into molybdate reactive (assumed to be principally inorganic; WEPi) and molybdate unreactive P (assumed to be principally organic; WEPo). Water extractable Pi concentrations showed no significant differences between soils from different land use classes (p = 0.97). The mean WEPi for extensive soils was skewed by the high concentration value for the Dartmoor soil (31 mg P kg−1). WEPo differed significantly between land use classes (p b 0.001), with similar values between arable and intensive grassland being significantly smaller than values for extensive soils especially the large concentrations found for Dartmoor and Southlakes South moorland soils. 3.2. Concentrations and distributions of P species amongst land use classes Neither total soil P determined by NaOH fusion nor the total P extracted by the NaOH–EDTA differed significantly between land use classes (p = 0.57 and 0.26, respectively; Table 1). The NaOH–EDTA matrix recovered an overall 72 ± 4% of the total soil P but some low recoveries suggested the presence of a pool termed ‘unextracted, alkaline insoluble P’ being N50% total P in five soils (Table S3). Concentrations of unextracted, P did not significantly differ between land use classes (p = 0.31) but were generally smaller in intensive grassland soils. The concentrations of P species in the alkaline soil extracts determined by 31 P NMR are plotted in Fig. 2 including comparison with literature global 31 P NMR studies (studies are detailed in Supplementary Information Table S5). Differences between land use classes were significant for ortho-Pi (p = 0.001; Table 1) but were not significant for concentrations of P-mono (p = 0.07). Due to The proportions of inorganic forms (ortho-Pi) and organic species varied greatly between land use classes (Fig. 2). Mean inorganic orthophosphate P (Ortho Pi) concentrations decreased in the order: arable (695 ± 328 mg P kg−1) N intensive grassland (462 ± 120 mg P kg− 1) N extensive land use (188 ± 82 mg P kg−1). However, arable soils had lower concentrations and proportions of Po. Concentrations of total monoesters (P-mono) followed the order intensive grassland (401 ± 94 mg P kg−1) N extensive (353 ± 173 mg P kg− 1) N arable (243 ± 66 mg P kg−1). For a number of other P species the effects of land use on concentrations could not be tested due to strongly skewed populations (associated with many

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M.I. Stutter et al. / Geoderma xxx (2015) xxx–xxx


Table 1 Means for soil properties, P indexes and P species concentrations according to land use classes with overall differences expressed by ANOVA. Different letters for classes denote significant differences according to Tukey tests (p b 0.05). Values are as mg kg−1 soil, except for pH, Psat, the % contributions of ortho-Pi and P-mono (compared to total P) and where stated.

pHCaCl2 SOC (g kg−1)1 WEOC


Alox (g kg−1) Feox (g kg−1) Soil total P by NaOH fusion Psat


Olsen P1 Fe–P1 Citric


Microb P1 WEPi1 WEPo1 P-NaOH–EDTA1 Ortho-Pi1 P-mono P-di P-polyp-end P-poly-mid P-phosphon % ortho-Pi % P-mono 1 2


Arable (n = 13)

Arable buffer (n = 3)

Intensive grassland (n = 10)

Extensive (n = 6)

Overall ANOVA2

5.95a (5.17–7.05) 22.4b (11.5–35.1) 0.13b (0.08–0.17) 2.2a (0.9–4.9) 5.3a (1.9–8.4) 1327a (650–3435) 0.17a (0.04–0.28) 69.4a (25.4–184.4) 45.8a (8.3–139.0) 20.9a (0.0–86.1) 1.9b (0–11.7) 3.9a (0.5–12.2) 1.5c (0.4–5.0) 938a (427–2966) 695a (276–2520) 243a (105–446) 0 (0–0) 1 (0–8) 0 (0) 0 (0) 55a (24–N100) 17b (7–28)

5.12ab (4.03–5.82) 54.0ab (20.5–78.3) 0.29ab (0.18–0.35) 1.9a (1.0–2.6) 5.7a (2.6–7.8) 894a (615–1223) 0.11ab (0.04–0.10) 41.5ab (27.7–50.7) 26.2a (5.7–51.9) 7.6a (2.4–12.3) 3.0ab (0.8–7.4) 3.9a (1.5–6.0) 2.0bc (0.9–3.7) 605a (436–712) 356ab (178–551) 230a (162–309) 7 (0–20) 11 (0–18) 0 (0) 0 (0) 38ab (29–45) 21ab (11–27)

5.22ab (4.20–6.56) 51.1a (20.6–100.7) 0.41a (0.16–0.91) 2.4a (0.7–4.0) 8.3a (3.8–12.0) 1185a (886–1865) 0.11ab (0.04–0.14) 43.3ab (9.2–86.8) 25.0a (8.0–49.2) 12.2a (2.5–27.0) 6.2ab (0.1–23.3) 3.5a (1.6–9.1) 5.3ab (1.4–11.2) 900a (596–1524) 462a (233–842) 401a (200–658) 9 (0–50) 29 (1–78) 0 (0) 0 (0) 39a (18–56) 26a (17–36)

4.79c (3.28–6.57) 158.0a (21.4–448.6) 0.75a (0.22–1.40) 2.3a (0.4–4.5) 6.9a (0.8–13.9) 1058a (294–1431) 0.06b (0.03–0.07) 18.5b (5.6–28.4) 31.7a (6.3–108.9) 15.6a (3.2–42.0) 15.9a (0.9–63.4) 7.5a (1.0–31.5) 17.5a (2.2–61.4) 660a (104–1062) 188b (67–334) 353a (37–621) 50 (0–102) 43 (0–100) 8 (0–50) 17 (0–80) 19b (9–27) 26ab (13–43)

* *** *** ns ns ns *** *** ns ns * ns *** ns *** ns nd nd nd nd *** *

Data were log10 (n, or n + 1) transformed where necessary according to Andersen–Darling testing (p N 0.05). ns, *, **, *** denote not significant and significance levels at p b 0.05, 0.01 and 0.001 levels, respectively; nd denotes that ANOVA was not determined as data could not be normalised.

samples below 31P NMR analytical detection limits). Examples of this were for P-polyp-mid and P-phosphon where concentrations of up to 50 and 80 mg P kg − 1, respectively, were only detected in two soils with extensive land use. The concentrations of diesters (P -di ) and polyphosphate end groups (P-polyp-end ) could not be tested since normality could not be satisfied even after transformations. In general though concentrationsfollowed the order extensive (50 ± 32 mg P kg− 1)− N intensive grassland (9 ± 10 mg P kg−1)− N arable (below detection) for P-di, and extensive (43 ± 28 mg P kg−1) N intensive intensive grassland (29 ± 13 mg P kg−1)− N arable (1 ± 2 mg P kg−1) for P-polyp-end. The percentage contributions of P species are considered compared to the total soil P by NaOH fusion and therefore only sum to 100% when the ‘unextracted, alkaline insoluble P’ is considered. When expressed in terms of % contributions the %ortho-Pi and additionally the P-mono showed significant differences between land use classes (p = b0.001 and 0.03, respectively; Table 1). Of the total soil P arable soils had a dominance of ortho-Pi (with 50% mean and range 24– 91%) N P-mono (20%, 9–35%) ≫ P-di (0%), P-polyp-end (0%) compounds, with mean 30% (0–67%) of unextracted P. Intensive grassland had ortho-Pi (39%, 18–56%) N P-mono (33%, 22–46%) N P-di (1%, 0–4%), P-polyp-end (2%, 0–6%) forms, with mean 25% (15–34%) of unextracted

P. Extensive semi-natural systems were dominated by monoesters (P31%, 13–50%) N ortho-Pi (19%, 9–27%) ≫ P-di (4%, 0–7%) and Ppolyp-end (3%, 0–7%). Diester P was present in none of the arable soils, one arable buffer soil and between 2 and 4% of total soil P in three intensive grassland soils (Southlake north, Canol and Bronydd low P). However, diester P commonly occurred in all the extensive soils (2–7% total soil P), although was absent from the forest soil. Polyphosphates were almost exclusively P-polyp-end (− 4 ppm). Only the Dartmoor Histosol contained a shift for P-polyp-mid at − 20 ppm indicative of long-chain polyphosphates (3% of total P). Phosphonates were only present in the extensive use Vealand unimproved (2% of total P) and Dartmoor soils (6% of total P). These species have a direct C–P bond and likely represent 2-aminoethylphosphonic acid at 20.6 ppm and phosphonoethanolamine at 16.5 ppm (Turner et al., 2007), the latter occurring only in the Dartmoor Histosol. For the Tentsmuir forest soil the NaOH–EDTA recovery of total soil P was low (35%) but the P species observed in the extract were dominated by ortho-Pi and P-mono. A weak shift at ~6.8 ppm present across all soil land use classes (12 soils in total up to 2% of P) has previously been suggested by Turner et al. (2007) as an unspecified stereoisomer of phytate. The strong shift at 4.2 ppm observed in all but two of the soils was attributed to the presence of scyllo forms of phytate accounting for 4% (0–7%), 7% (5–10%) mono;

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P determined by P NMR (mg P kg )



P determined by P NMR (mg P kg )

P determined by



P NMR (mg P kg )


2600 2400 1400

(a) Arable soils

1200 1000 800 600 400 200 0 4500 4000 1400 1200

(b) Intensive grassland soils

1000 800 600 400 200 0 800

(c) Seminatural soils







Inorganic Orthophosphate



Orthophosphate monoesters



Orthophosphate diesters

Fig. 2. Concentrations of P species in UK soils determined by 31P NMR, compared to global soil data. Box (25 and 75 percentiles with median marked) and whisker (5 to 95% percentiles) plots of concentrations (mg P kg−1 dry soil) of the three main classes of P compounds (inorganic orthophosphate, orthophosphate monoesters and orthophosphate diesters) for grouped agronomic management systems: (a) arable (UK n = 13, global n = 38), (b) intensive-grassland (UK n = 10, global n = 11) and (c) extensive/semi-natural (UK n = 6, global n = 20). In each case, UK data from the current study is on the left of each pairing (to be coloured red in final figure) and global soil literature values to the right (grey boxes). The sources of the global data are presented in Supporting Information Table S5. (For interpretation of the references to color in this figure legend, the reader is referred to the web version of this article.)

and 4% (0–7%) for arable, intensive grassland and extensive land classes, respectively. The ratio of monoesters to scyllo form monoesters was not significantly different between land class being 4.0 (3.3–5.2), 4.0 (2.9– 5.9) and 5.5 (4.0–6.8) for arable, intensive grassland and extensive, respectively. Expansion of the fine structure of the monoester region is depicted for the representative spectra in Fig. 1. In the absence of a

procedure for accounting for losses of RNA and certain phospholipids sensitive to degradation in the alkaline extracts the P-di detected here as a broad signal ~−0.2 ppm was likely to be DNA since this is known to be stable in alkaline extracts (Turner et al., 2007). For arable and extensive semi-natural soils the concentrations of the three principal P forms (ortho-Pi, P-mono and P-di) were similar between

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M.I. Stutter et al. / Geoderma xxx (2015) xxx–xxx Table 2 Correlation matrix (Pearson's r) of concentrations of the principal 31P NMR defined P species Inorganic orthophosphate (ortho-Pi) and monoesters (P-mono) against soil properties and standard P indices (n = 32).

pHCaCl2 WEOC1 SOC1 Alox Feox Olsen P1 Fe–P1 Psat1 Po_citric1 Microb P1 WEPi1 WEPo1



ns ns ns 0.46⁎⁎ 0.63⁎⁎⁎ 0.65⁎⁎⁎ 0.55⁎⁎⁎ 0.37⁎ 0.44⁎

ns 0.44⁎⁎ 0.44⁎ 0.34⁎ 0.56⁎⁎⁎ ns ns ns 0.36⁎ 0.40⁎⁎ ns 0.50⁎⁎

ns ns ns

ns, *, **, *** denote not significant and significance levels at p b 0.05, 0.01 and 0.001 levels, respectively. 1 Transformed by log10 (n, or n + 1) according to Andersen–Darling test results (p N 0.05).

our UK data and the global data taken from the literature (Fig. 2), showing stability in P compound accumulation across broad global geoclimatic regions. In contrast, intensive grassland soils had considerably greater variability in the P forms globally than in UK soils.


trends in relation to the position of data points. There was a clear split of the site data points between the negative and positive sides of PC1. Arable sites were exclusively negative on PC1 but also in this region were the arable buffer sites, 2 intensive grassland (Bicton, De Bathe both low in SOC) and the Tentsmuir forest site and corresponded to the angles and positions of the loadings of Psat, Olsen P and ortho Pi). The remaining intensive grassland sites and a further extensive site (Dunsdon unimproved) plotted between 0 and 2 on PC1, whilst the remainder of the extensive/semi-natural sites plotted more positively on PC1 and corresponded to angles and positions for loadings of Feox, WEOC, P-poly-end, WEPo, P-di, SOC, Microb P, P-mono and Alox. In addition (data not shown) PC3 accounted for an additional 14% of variance (cumulatively 74%, with further components added much less), with loadings dominated by Feox (0.55), P-mono (0.42) and WEPi (− 0.36). The significance of differences in PC1, PC2 and PC3 by one-way ANOVA between the land use classes was p = b0.001, 0.59 and 0.38, respectively. Ratios of WEPo/WEPi were 0.5 ± 0.2 for arable, 2.1 ± 0.8 for intensive grassland and 6.0 ± 6.5 for extensive and indicated that soluble Po dominated over Pi in all but arable soils. Table 2 and Fig. 3 show that WEPi was not closely related to any individual parameters or multivariate groups. Conversely, WEPo was closely related to P-polyp-end, P-di and P-mono and highly significantly correlated with all of these (p b 0.001). 3.4. Soil C and P relationships

3.3. Relationships between soil properties and distribution of P forms

Ortho Pi





Olsen P Psat

Al ox P-mono Microb-P SOC P-di WEPo


P-poly-end WEOC

Soils were purposely sampled to maximise their range of soil P and C concentrations. Fig. 4 plots in log–log space the relationship between the Po/Pi balance (where Po is orthophosphate monoesters and Pi inorganic orthophosphate in the alkaline extracts) against SOC concentrations. The relationship is shown for soils of the present study according to land use, with a positive relationship extending from arable soils in the area of ortho Pi dominance and low SOC content (bottom left) through grassland to semi-natural soils with a dominance of Po and large SOC contents (top right). The mean (±1 S.E.) of the Po/Pi ratios

Ratio monoester / inorganic ortho-P

Concentrations of ortho-Pi (Table 2) showed significant positive correlations with a number of common P indices; Alox, Feox, Olsen P, Fe–P and Psat and P-o-citric. Monoester P concentrations had significant positive correlations with WEOC, SOC, Feox, Alox, P-o_citric, Microb P and WEPo. P-di and P-polyp-end were highly significantly positively correlated with WEOC, SOC, Microb P and WEPo, but only P-polyp-end was weakly positively correlated with the stabilising Feox surfaces across all soils. Fig. 3 shows the first two principal components of the combined data (together accounting for 60% of the variance. Loadings (proportional to the distance of the arrow heads along lines projecting from the origin) for PC1 are dominated by P-di (0.37), P-polyp-end (0.35), SOC (0.36), WEOC (0.36), and Microb P (0.31) and for PC2 loadings are dominated by Fe–P (0.49), Olsen P (0.42), ortho-Pi (0.40) and Psat (0.38). Common angles of variables from the origin show similar explanatory



0.1 1



SOC (g





Arable buffer

Intensive grassland


Global soils Fig. 3. Principal components biplot of soil properties, soil P indices and 31P NMR P forms. Arrow heads along projected lines (dotted for general soil properties, bold for NMR P forms) show the direction of loadings for each variable, the separation of data points (denoted according to land use classes in Table 1) and corresponding positions relative to arrow heads shows the importance of the different analytical variables.

Fig. 4. Relationships between the distribution of soil P between monoester and inorganic forms according to C contents. Soil monoester P:inorganic orthophosphate ratios are plotted against soil organic C contents for soils of the present study (in log–log space) and these compared with global soils data (see Supporting Information Table S5 for data sources).

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was 1.77 ± 0.28 for extensive, N 1.01 ± 0.18 for intensive grassland, N0.81 ± 0.28 for arable buffer and N 0.42 ± 0.05 for arable soils. In comparison the global soils data (Supplementary Information Table S5) have considerably greater scatter, likely related to the large range in climates compared to the temperate UK soils of the present study. Data for tropical soils (Ethiopia, Columbia, Thailand) represent the greatest separation with high Po/Pi ratios at low to moderate SOC contents. Organic C:P ratios have previously been consider as providing approximate thresholds for microbes to tend towards either net P mineralization (SOC:P b 200) or net P immobilization (SOC:P N 300) (Dalal, 1977). For the present study this ratio used the SOC concentration divided by the sum concentration of organic P species determined in the alkaline extracts. The organic C:P ratios (means ± 1 S.E.) were 379 ± 114 for extensive soils, N 209 ± 51 for arable buffer soils, N 115 ± 19 for arable and similar to 117 ± 10 for intensive grassland soils. The highest ratios occurred for the extensive land class Dartmoor and King's Sedgemoor Histosols but also for the forest site, which had small SOC and very small Po concentrations. There was no significant relationship across all sites between Po/Pi and organic C:P ratios but the means for the four land use classes showed a general positive relationship. This indicated that P-mono accumulated relative to Pi at greater organic C:P ratios (ie ratios tending to favour immobilization). However, intensive grassland deviated tending towards lower C:P ratios for a given Po/Pi ratio (ie maintaining P-mono accumulation at C:P ratios favouring mineralization). 4. Discussion 4.1. Soil P forms under different land uses These 32 soils from the UK were selected on the basis of land use and to maximise variation in P and C contents to evaluate how species and concentrations of P in soils will vary according to management and soil property interactions. Our primary hypothesis was that land use and inherent management was a key control on the accumulation of quantities and forms of soil P. However, due to the considerable concentration variations within land use classes there were no significant differences (p N 0.05) between land use classes for the concentrations of total soil P monoester organic forms (Table 1). However, ortho-Pi had significantly smaller concentrations and relative contribution to total P in extensive/semi-natural soils than intensive grassland and arable soils. Additionally, when P-mono contributions to total P were considered intensive grassland had greater %P-mono than arable. In general differences were observed in terms of the presence of more labile organic P forms in some intensively grazed and semi-natural (moorland, rough grassland) soils, but absent in arable soils. Therefore aspects of land use and management are discussed below as influential but there was a substantial relationship with key soil properties both between and within land use classes. Hence a complex interplay of factors controlled soil P accumulation and the P species occurring. The dominance of Pi in arable soils (and especially the higher values in the range 276 to 2520 mg kg−1) and significant correlations (p b 0.001) with Olsen P, Fe–P and Psat (Table 2) suggests a history of large fertiliser inputs. Although we do not have evidence for these inputs and they are assumed from typical UK management regimes it is likely that chemical fertiliser input rates strongly influence soil Pi contents within the arable class. UK arable soils had lower concentrations of P-mono and P-polyp-end than soils under other management and non-detectible levels of P-di, P-phosphon and P-poly-mid and this matched P species distributions from global literature 31P NMR studies (38 arable soils detailed in Supplementary Information Table S5). In soils under intensive grassland, Po and Pi concentrations were equivalent. The dominance of P-mono in grasslands compared to other cultivated soils has been observed previously (Harrison, 1987; Turner et al., 2003b). However, Turner et al. (2003b) found smaller P-mono concentrations than in the present study. Whilst cultivation and crop

removal limits Po accumulation in arable soils (Solomon and Lehmann, 2000) between 60 and 95% of plant utilised P in grazed grasslands is returned to the soil in plant litter and animal excreta (Condron et al., 2005), although manures may contain ortho-Pi or Pi that degrades to Po over time. Using 31P NMR Turner et al. (2003b) found concentrations of 74 to 561 mg P kg− 1 ortho Pi, 154 to 751 mg P kg− 1 P-mono and 11 to 109 mg P kg− 1 as P-di in 29 lowland permanent pasture soils in England and Wales. In our study the Banadl grassland soil had P concentrations at or exceeding the top of the ranges found by Turner et al. (2003b) with 842 and 658 mg P kg− 1 as ortho-Pi and total P-mono, respectively. Although not testable by ANOVA due to numerous occurrences of values below detection limits strong general differences were shown between the presence and concentration of P-di and P-polyp-end species between land uses (Table 1). However, P-di concentrations in the present study were considerably more variable than reported by Turner et al. (2003b). Compared to other studies the extensive grazing systems at Dartmoor and Southlake South showed large diester concentrations (thought mainly to be DNA) even though we did not apply correction factors for diester degradation during extractions. Conversely, P-di was not detected in seven out of ten intensive grassland soils, two of three arable buffer soils and all the arable soils. The intensive grassland soils analysed in the present study showed less variation in P form concentrations relative to the summarised values from global literature (Fig. 2). This discrepancy could arise from a narrow range of stocking rates, animal types, fertiliser input types and rates and/or climate in the UK. The intensive grassland soils were all sampled in a distinct climatic region of the UK (SW of England; mild and wet) where intensive grassland dominates. In contrast the concentration ranges of P species for UK extensive land use soils matched distributions of global soils across differing climates.

4.2. Prominence and stabilisation of soil P forms We observed that microbial P cycling indicators (P-di, P-polyp-end, Ppositively related to the presence of SOC and WEOC (i.e., total and labile SOC) (Table 2, Fig. 3). It is widely accepted that plants uptake P as ortho-Pi and that phosphate from organic sources must first be hydrolysed. The resupply of ortho-Pi occurs via desorption from soil surfaces, dissolution of phosphate minerals, and mineralization of Po mediated by extracellular specific and broad acting phosphatase enzymes to hydrolyse P esters (Frossard et al., 2000). In soil, microbes and plants take up ortho-Pi during growth, returning Po to soils as lysed cell contents, plant detritus, and excreta in grazed systems. Greater concentrations of P-di and P-polyp-end in grassland and extensive soils (Table 1) are indicative of greater biological P turnover (Turner et al., 2003b; Frossard et al., 2000) and an active microbial population efficient in turnover of soil Po in conditions of limited readily available inorganic P (Cross and Schlesinger, 1995; Bünemann et al., 2012). Diesters comprise nucleic, techoic acid and phospholipids microbial constituents (Condron et al., 2005). However, current understanding is that the alkaline extraction for NMR leads to degradation of certain sensitive diester forms (Turner et al., 2007). Polyphosphates are another indicator of microbial P cycling (Turner et al., 2003a), thought to be linked to microbial P storage during times of ‘luxury’ supply or environmental stress (Condron et al., 2005) and comprise adenosine di- and tri-phosphates from bacteria, protozoa, fungi, algae and insects (Condron et al., 2005). Turner et al. (2003b) showed that P-di concentrations in UK grasslands were positively related to microbial biomass, confirming work by Makarov et al. (2002). In the intensively managed soils the absence of peaks for DNA components of P-di in our extracts and polyphosphates indicates the loss of the natural soil–plant-microbial ability to recycle organic P in soils reliant on fertiliser inputs (Condron et al., 1990). This is exacerbated in arable soils where crop monocultures reduce the microbial diversity (Stutter and Richards, 2012) and crops have been cultured for rapid uptake of inorganic P. phosphon)

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Balances between different organic vs inorganic P compounds are influenced by sorption interactions with soil surfaces (Stewart and Tiessen, 1987). Certain monoesters (e.g. myoIHP, although not specifically identified here) are charge-dense and adsorb strongly by ligand exchange to soil surface complexes that bind and stabilise ortho Pi (Celi et al, 1999) and additionally SOC. Monoester P concentrations were strongly correlated with soil amorphous Fe (and to a lesser extent Al) phases (Table 2; as oxalate extractable Feox and Alox). Sorption to these reactive surfaces imparts protection to labile P forms against microbial decomposition, or enzyme hydrolysis (Giaveno et al., 2010).


The considerable P-mono concentrations accumulated in agricultural topsoils present opportunities and challenges to crop and soil science to unlock soil P reserves and increase future efficiencies of agronomic systems (Stutter et al., 2012). In comparison, sorption of P-di (with one ionisable proton per molecule) is considerably weaker than for P-mono, leaving P-di susceptible to microbial decomposition and enzymatic attack. The highly significant correlations of P-di with SOC and WEOC support the stabilisation of diesters with humic acids, as proposed by Makarov et al. (2002). Lowering soil pH increases the sorption of diesters and phosphonates (Condron et al., 2005; Turner and Blackwell,

Agronomic and site factors : Chemical fertiliser inputs, loss of plant residues through cropping. Freely draining, low C soils.

Manure and chemical fertiliser. Soil C accumulates under more permanent grass cover. Range of soil conditions.

Limited fertiliser inputs, high soil moisture and low temperature increasinglyconstrains agriculture.

Organic C:P 0





Immobilization Mineralization

P form distributions: Arable





natural soil

Soil C and P biogeochemistry Labile organic C and P suffers burn out by rapid organic matter turnover. Soil Pi dominates and maintains P availability to a range of crops. Ecosystem microbial functioning is impaired

Microbially-mediated P immobilization increasingly requires labile C sources Increasing microbial cycling due to litter inputs, increased soil C and manure inputs increase microbial cycling. As a result the diversity of labile Pspecies increases. Lowering pH favours sorption and stabilization of diesters, phosphonates and polyphosphates

Sorption to mineral phases protects against microbial and enzyme decomposition. Variability in soil Fe, Al complexes, other stabilizing factors (pH, clay) and microbial communities governs variability in specific soil response.

P form distributions (clockwise from top and excluding ‘unconfirmed P’): Inorganic P





Fig. 5. Conceptual diagram of linkages between the biogeochemistry of soil C and P across different land uses and soil properties.

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2013) explaining the increased prevalence of these labile P forms in wetter, humified acid soils and particularly phosphonate concentrations in the Dartmoor moorland soil. 4.3. Soil C as a driver of P turnover Alongside abiotic factors of stabilisation the microbial processing of P controls the Po/Pi balances and overall soil P accumulation. Table 2 shows that Microb P, SOC and WEOC all significantly positively correlated with concentrations of organic P forms (P-mono, P-di and P-polyp-end) that were more prevalent in soils that were grazed or set aside than in arable soils. Our soil sampling design maximised the range of soil C:P concentrations to evaluate relationships between soil C and P species distributions. However, the limitation of this was an unbalanced sample design (see Supplementary Information Table S1) where Histosols are only present in the extensive land use group (ie it was not possible to find high C soils with high Pi inputs). This is due to the moisture (and other climatic and topographic factors) placing constraints on such soil types supporting arable and more intensively managed grasslands. Despite this Fig. 4 suggests that in temperate systems there is a tendency for accumulation of organic monoester P species relative to inorganic orthophosphates as SOC concentrations increase and that this generally holds true at a global scale. Soil organic C:P values of b200 and N300 have previously been used, respectively, to indicate the prevalence of mineralization vs immobilization of soil P (Dalal, 1977; Condron et al., 2005; McDowell and Stewart, 2006). In the present study the organic C:P threshold indicating a tendency for immobilization was only exceeded in extensive/seminatural soils (especially high C Histosols), whereas ratios indicative of mineralization occurred similarly for arable and intensively-grazed systems. Often these boundaries have been applied to understand the fate of organic matter applied to soils, but they have also been used to determine the relative balance of soil internal P cycling (McDowell and Stewart, 2006). Immobilization results from microbial Pi utilisation during the decomposition of organic residues with high C:P values and leads to conversion to Po in microbial products and plant litter (Condron et al., 2005). It could be viewed that decreasing Po/Pi and C/P are both a function of increasing inorganic fertiliser inputs associated with more intensive land use and that this is inseparable (given the data here) from inherent climate constraints governing land capabilities and P inputs. Indeed there were significant negative correlations of both C:P (r = − 0.61) and Po/Pi (r = −0.64) with soil Olsen-P. Bünemann et al. (2012) recently showed that under low Pi input conditions, immobilization dominates gross P fluxes in P depleted grassland soils and leading to the large Po/Pi ratios for extensive grassland soils. Conversely, the dominance of low C:P ratios for all arable and intensive grassland soils suggests long-term mineralization to supply Pi, especially for cropping systems where crops inorganic P demands are great and organic P returns to soils are interrupted by cropping. Malik et al. (2012) showed that inorganic fertiliser inputs tended to lead to Pi sequestration, where manure and inputs of other C rich fertilisers led to both labile and sequestered Po accumulation. However, the controls of system C inputs on P turnover are undoubtedly complex. Kirkby et al. (2013) demonstrated that inorganic N and P availability was required to sequester plant residue C into stable soil pools, and in turn the P was sequestered alongside C. Clearly there are many inter-related abiotic and biotic factors governing accumulation of the P species in these soils (as in Fig. 3). For example the presence and persistence of both C and P species may be strongly influenced by protection from enzyme hydrolysis afforded by sorption to soil surfaces (often termed ‘mineral protection’). Ohno et al. (2011) recently showed that the presence of orthophosphate, polyphosphate and diester DNA compounds in soils strongly related to oxalate-extractable Fe and Al surfaces. This observation may confirm ‘mineral protection’ of labile Po species directly, or via soil organic

matter interactions with these surface complexes (see correlations, Table 2). Charge dense forms of P-mono may overcome competition for sorption sites against DOC anions, where other Po components remain more soluble and readily decomposed. This may be conceptualised as a microbial ‘push’ for Po storage and a geochemical ‘pull’ for storage via stabilisation. These factors are summarised in a conceptual model in Fig. 5. 5. Conclusions Using a range of 32 soils from the UK our study has provided a detailed picture of the links between P concentrations and species abundance in soils, land use and soil properties. Arable soils are shown to be dominated by ortho-Pi with P-mono species associated with strongly sorbing Al and Fe soil surfaces. Orthophosphate monoesters dominate in intensive grasslands. In contrast concentrations of labile P-di, P-polypend (and in the case of moorland soils P-polyp-mid and P-phosphon) species increase as grazing becomes extensive, associated with labile soil organic matter and microbial turnover. Hence the diversity of P species in the less intensively managed soils and the statistical links made in this study to microbial P, total and labile pools of soil C can be viewed as positive indicators of ecosystem function and diversity. The accumulation of the different forms of soil P is related to land uses as we hypothesised but this is complicated by the large variation between soil, site and management factors within classes of land use. Fertiliser inputs (not determined here) may largely explain variation in Pi concentrations and prevalence over Po forms but numerous soil P stabilisation and microbial processing factors interact to govern P turnover and accumulation. Amongst the soil factors studied here reactive properties of SOC and analytically-defined properties oxalate extractable Al and Fe seem to be particularly important. Future studies should seek to build on this knowledge by incorporating these factors in soil P studies alongside evidence for fertiliser input history and process-indicators such as discrimination of labile forms of soil C. Improved understanding of such factors is needed to design agricultural systems (in terms of where to target, nature of crops and input rates) to push for greater yields per input and with less negative environmental consequences such as P losses to waters. Acknowledgements We thank the Scottish Government Rural Environment Research and Analysis Directorate and UK Biotechnology and Biological Sciences Research Council (BBSRC BB/K018167/1) for funding aspects of this research. We gratefully acknowledge the assistance of C. Taylor, H. Taylor, S. Richards and R. Wendler in soil collection and analysis. Appendix A. Supplementary data Supplementary data to this article can be found online at http://dx. References Bedrock, C.N., Cheshire, M.V., Chudek, J.A., Goodman, B.A., Shand, C.A., 1994. 31P NMR studies of humic acid from a blanket peat. Humic Substances in the Global Environment and Implications on Human Health, pp. 227–232. Bradford, M.A., Fierer, N., Reynolds, J.F., 2008. Soil carbon stocks in experimental mesocosms are dependent on the rate of labile carbon, nitrogen and phosphorus inputs to soils. Funct. Ecol. 22, 964–974. Bünemann, E.K., Bossio, D.A., Smithson, P.C., Frossard, E., Oberson, A., 2004. Microbial community composition and substrate use in a highly weathered soil as affected by crop rotation and P fertilization. Soil Biol. Biochem. 36, 889–901. Bünemann, E.K., Oberson, A., Liebisch, F., Keller, F., Annaheim, K.E., Huguenin-Elie, O., Frossard, E., 2012. Rapid microbial phosphorus immobilization dominates gross phosphorus fluxes in a grassland soil with low inorganic phosphorus availability. Soil Biol. Biochem. 51, 84–95. Cade-Menun, B.J., Liu, C.W., 2014. Solution 31P-NMR spectroscopy of soils from 2005– 2013: a review of sample preparation and experimental parameters. Soil Sci. Soc. Am. J. 78, 19–37.

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