adsorption controls mobilization of colloids and leaching of dissolved phosphorus

Adsorption controls mobilization of colloids andleaching of dissolved phosphorus

J . SIEMENS, K. ILG, F. LANG & M. KAUPENJOHANN

Department of Soil Science, Institute of Ecology, Berlin University of Technology, Salzufer 11–12, 10587 Berlin, Germany

Summary

Loss of phosphorus (P) from agriculture contributes to the eutrophication of surface waters. We have

assessed the magnitude and controls of P leaching and the risk of colloid-facilitated transport of P from

sandy soils in Munster. Concentrations of soluble reactive P in drainage water and groundwater were

monitored from 0.9 to 35m depth. Total P concentrations, P saturation, and P sorption isotherms of soil

samples were determined. Concentrations of dispersible soil P and colloidal P in drainage water and

groundwater were investigated. The concentrations of soluble reactive P in drainage water and ground-

water were close to background concentrations (< 20�g P l�1). Median concentrations in excess of

100�g P l�1 were found down to 5.6m depth at one of four research sites and in the lower part of the

aquifer. Experimentally determined equilibrium concentrations and the degree of P saturation were good

predictors of P concentrations of drainage water. Large concentrations of dispersible P were released

from soil with large concentrations of oxalate-extractable P and addition of P induced further dispersion.

Colloidal P was transported in a P-rich subsoil when there was a large flow of water and after nitrate had

been flushed from the soil profile and total solute concentrations were small. We conclude that the

concentration of soluble reactive P in drainage water is controlled by rapid adsorption in the sandy soils.

Subsurface transport of dissolved P contributes substantially to the loss of P from the soils we investi-

gated. Accumulation of P in soils increases the risk of colloid-facilitated leaching of P.

Introduction

Eutrophication of surface waters is often caused by input of

phosphorus (P). It has been estimated that agriculture con-

tributes 28% to the P loading of surface waters in Germany

(Werner, 1997). Because inorganic P is strongly sorbed by

soils, surface runoff and erosion have been regarded as the

most important vectors of P loss from agricultural land to

surface waters (e.g. Sharpley et al., 1994). However, the leach-

ing of P from soils to groundwater and drains has received

increasing attention (e.g. Breeuwsma & Silva, 1992), and

Heckrath et al. (1995) and McDowell et al. (2002) have

observed that the leaching of P increases as a threshold of

the P sorbed to soils is exceeded.

Munsterland is densely stocked with cattle and pigs, which

might lead to an accumulation of P in soils (e.g. Leinweber

et al., 1997). Furthermore, Plaggic Anthrosols, containing

large P concentrations (> 110mgPkg�1 in citric acid,

Miedema, 1991) and developed on Quaternary sands, are

common. Thus, P leached from the soil is likely to reach the

groundwater with unfavourable consequences for the environ-

ment.

Colloidal P might be lost from agricultural soils by leaching.

The association of strongly sorbing compounds with colloids is

thought to enhance mobility (e.g. Kretzschmar et al., 1999).

Indeed, P is bound to colloids (Haygarth et al., 1997; Hens &

Merckx, 2001), and there are indications that suspended par-

ticles and colloids act as carriers (Laubel et al., 1999). Because

sorption of P to Fe oxides increases colloid stability (Puls &

Powell, 1992), Plaggic Anthrosols rich in P are susceptible to

the leaching of colloidal P.

Our objectives were (i) to quantify the leaching of dissolved

P and to identify its controls and (ii) to evaluate the risk of

mobilization and transport of colloidal P in Munster.

Materials and methods

Sites and agricultural management

We investigated four research sites in the ‘Munsterland

Kiessandzug’ geological unit near Munster, northwest Germany

(Staude, 1986). The mean annual precipitation is 742mm,Correspondence: J. Siemens. E-mail: [email protected]

Received 8 October 2002; revised version accepted 21 July 2003

European Journal of Soil Science, June 2004, 55, 253–263 doi: 10.1046/j.1365-2389.2004.00596.x

# 2004 Blackwell Publishing Ltd 253

and the mean annual temperature is 9.3�C. The soils at sites A,

D and H are characterized by a 60–70 cm thick anthropogenic

layer, which is enriched in organic C (Table 1). They are

classified as Plaggic Anthrosols in the FAO scheme. The soil of

site S is a Gleyic Podzol. The soils of sites S and H are developed

from the Saalian glacial sands of the Kiessandzug, those at sites

A and D are in Weichselian aeolian sands. The groundwater

table is at 4–5m at site H. It can rise to 0.7m below the soil

surface during spring at site S. Perched water tables occur below

1.4m depth at site A and below 2.2m depth at site D because of a

dense layer of silt (‘Schlufffolge’, Staude, 1986). Siemens et al.

(2003) show the location of the research sites and the spatial

distribution of soils and groundwater conditions.

In 1993, sites S and D were converted from arable land to

permanent fallow. Since then they have received no fertilizer

and the grass that was grown there remained on the sites. Sites

A and H were used for a crop rotation of ryegrass (Lolium

perenne), maize (Zea mays), winter barley (Hordeum vulgare),

triticale (a Triticum–Secale hybrid) and other cereals. From

August 1999 until May 2001, 54 kg P ha�1 was applied at site

A as manure, and 20 kgP ha�1 as mineral P. At site H,

104 kg P ha�1 as manure and 20 kg mineral P ha�1 were

applied from the seeding of ryegrass in March 1999 until the

end of the experimental period in May 2001. Again, Siemens

et al. (2003) give details on the agricultural management,

including rates and dates of fertilization.

Sampling and analysis of drainage water and groundwater

Drainage water was sampled with tensiometer-controlled suc-

tion plates (0.9–3.3m depth, five replicates per depth, Siemens

Table 1 Soil and sediment features of the experimental sites and instrumentation

Soil and sediment properties Instrumentation

Depth /m Horizon Texturea Corg /% pH in water Sampler typeb Number of replicates Depths /m

Site A: Plaggic Anthrosol, groundwater table: 3–4mc

0–0.4 Ap fs 1.11 5.8

0.4–0.7 A fs 0.12 5.7

0.7–1.9 Bg fs – ms 0.04 5.2 SP 5 0.9, 1.5

1.9–2.6 C1 si ND 5.5 SP 5 2.2

2.6–3.3 C2 ms ND 5.9 SC 2 2.8, 3.1

3.3–5.6 C3 si ND 7.5 SC 2 4.0, 5.6

Site H: Plaggic Anthrosol, groundwater table: 4–5m

0–0.4 Ap fs 1.27 5.8

0.4–0.6 A fs 0.40 5.8

0.6–0.7 Bw fs 0.15 5.2

0.7–1.3 CB ms 0.05 5.6 SP 5 0.9

1.3–2.7 C1 ms ND 5.4 SP 5 1.5

2.7–4.1 C2 fs ND 6.0 SP 5 3.3

4.1–5.0 C3 fs ND 7.1 SC 2 3.8, 4.4

Site D: Plaggic Anthrosol, groundwater table: 3–4mc

0–0.3 Ap fs 1.14 5.5

0.3–0.7 A fs 0.49 6.4

0.7–1.0 Bg1 fs – ms 0.10 6.2 SP 5 0.9

1.0–2.3 Bg2 ms 0.05 5.8 SP 5 1.5, 2.2

2.3–2.5 Bg3 cs 0.04 5.2

2.5–2.9 Bg4 si – fs 0.05 5.2 SC 2 2.8

2.9–3.6 Bg5 ms – cs 0.04 5.3 SC 2 3.1

3.6–4.4 Bg6 si 0.08 8.2 SC 2 4.0

Site S: Gleyic Podzol, groundwater table: 0.7–2m

0–0.3 Ap fs 2.51 6.1

0.3–0.7 Bhs fs – ms 0.73 5.8

0.7–1.1 Bg1 fs 0.13 6.0 SP 5 0.9

1.1–1.4 Bg2 fs 0.05 5.7 SC 2 1.3

1.4–2.7 C ms – cs 0.04 6.2 SC 2 1.8, 2.0

afs, fine sandy; ms, medium sandy; cs, coarse sandy; si, silty.bSP, suction plates; SC, suction cups.cPerched water tables during late winter and spring to 1.4m (site A) and 2.2m (site D); all depth denotations are depth below soil surface.

254 J. Siemens et al.

# 2004 Blackwell Publishing Ltd, European Journal of Soil Science, 55, 253–263

& Kaupenjohann, 2003) and suction cups (1.3–5.6m depth,

two replicates per depth; Table 1). Suction plates and suction

cups were made from borosilicate glass to minimize the sorp-

tion of P to porous cups and plates (Bottcher et al., 1984).

Samplers were equilibrated in the field at least 2weeks before

sampling started. We analysed samples that were collected

between November 1999 and May 2001. Groundwater samples

were obtained from two multilevel wells (ML7 and ML8).

Well ML7 is adjacent to site H, down the hydraulic gradient.

Well ML8 samples groundwater that originates from the

northern part of the catchment and is not influenced by drain-

age from the soils we investigated.

All samples were passed through PET filters (< 0.45�m

pore width, Macherey and Nagel, Duren, Germany) from a

PE syringe and stored at �18�C until analysis. We quantified

P (soluble reactive P, SRP) concentrations by the method of

Murphy & Riley (1962) using a Zeiss PM2K photometer

(Zeiss, Jena). The detection limit was 5�g P l�1.

Colloids range in size from approximately 10 nm to 1�m

(Kretzschmar et al., 1999), and so the soluble reactive P frac-

tion < 0.45�m may include colloidal P. Samples (n¼ 136)

from all sites and all depths that covered a large range of

SRP concentrations were ultra-centrifuged at 300 000 g at

10�C for 1 hour to remove colloids (Beckman Optima TL,

Unterschleissheim, Germany). Colloidal P in the soluble reac-

tive P was calculated as the difference between concentrations

of P in non-centrifuged and ultra-centrifuged samples. Add-

itionally, we determined colloidal P in solution samples taken

in February, April and May 2002. We tested the significance of

differences between P concentrations of ultra-centrifuged and

non-ultra-centrifuged samples by the Wilcoxon test.

The filter effect of the plates was tested by passing a suspen-

sion of 3.9mg l�1 and 1.4mg l�1 of colloidal P through the

plates. In the first case 50% of the colloidal P was removed by

the plates, in the second case this fraction was 29%. Thus, the

plates probably reduce the concentration of colloidal P during

sampling. It is unlikely, however, that no colloidal P passed

them at all.

We used a WTW LF90 conductometer (WTW, Weilheim) to

measure the electrical conductivity of the same set of samples

that was also ultra-centrifuged.

Concentrations of P in soils and degree of P saturation

Combined soil samples were taken during installation of suc-

tion plates and cups. We analysed samples from the top three

horizons of each site. These were Ap, plaggic horizon, and Bg

horizons for sites A and D, Ap, plaggic horizon and Bw for

site H and Ap, Bhs, and Bg1 for site S. For the sake of

simplicity we will denote the horizons of the different sites as

0–30 cm, 30–60 cm and 60–90 cm, which are approximately

their depths (Table 1). Soil samples were air-dried before

analysis. To determine total P contents, 10ml of concentrated

HNO3 (13M) was added to 0.5 g soil and heated to 185�C for

6 hours in a closed PTFE container. In addition to total P

contents, we determined oxalate-extractable P, Fe and Al

(Pox, Feox, Alox) as described by Schlichting et al. (1995,

p. 148) by extracting 2 g of soil with 100ml ammonium oxalate

(0.2M, pH3.25) for 1 hour in the dark. Iron and Al concentra-

tions were determined on a Perkin Elmer 1100B atomic

absorption spectrometer (Perkin Elmer 1100B, Shelton,

USA). We determined P by the method of Murphy & Riley

(1962) after diluting the oxalate extract to 200 times its

volume. We analysed two replicates of the combined samples

for the determination of total P concentrations and four repli-

cates for the determination of Pox, Feox and Alox.

The degree of saturation of the soil’s P sorption capacity

was characterized by the saturation index Z (Beek, 1978; van

der Zee & van Riemsdijk, 1986, 1988; van der Zee et al., 1988):

Z ¼ Pox½ �0:5 Feox½ � þ Alox½ �ð Þ ; ð1Þ

where [Pox], [Feox] and [Alox] denote the concentrations of

the elements in the oxalate extracts in mmol kg�1. The

denominator of Equation (1) approximates the total P sorp-

tion capacity of the soil (Beek, 1978; van der Zee & van

Riemsdijk, 1988). The total P sorption capacity accounts for

the rapid equilibrium adsorption–desorption reaction and a

slow, kinetic sorption reaction that is diffusion controlled

(van der Zee & van Riemsdijk, 1988). By doing long-term

desorption experiments, van der Zee et al. (1988) showed that

the fast, reversible sorption capacity is approximately

Z¼ 0.25. Large P concentrations >100�g l�1 may thus be

expected if Z �0.25.

Sorption experiments

Sorption isotherms were established by shaking 10 g of soil

with 20ml 0.01M KCl solution that contained 0, 0.07, 0.33,

1.63, 6.53, 16.32, 32.63, 48.95 and 65.26mg l�1 PO43–-P sup-

plied as KH2PO4 for 24 hours. The KCl solution was adjusted

to the soil pH determined in 0.01M CaCl2 solution. The

extracts were centrifuged at 3000 g for 10minutes and filtered

(Schleicher and Schuell, Dassel, Germany, No 512½, P-free

filter paper). The first 5ml of the filtrate was discarded. Sorp-

tion experiments were done in duplicate. Phosphate concentra-

tions were determined as described by Murphy & Riley (1962).

To account for P desorption from soil at small concentra-

tions of added P, we used a modified Langmuir equation to

describe the sorption isotherms:

s0 ¼ S0maxk

0c

1þ k0c� C; ð2Þ

where s0 is desorbed or sorbed P (mg kg�1), S0max is a

fitting parameter for the maximum of sorbed P (mg kg�1), k0 is

an affinity parameter (l mg�1), c is the concentration of dissolved

Adsorption controls colloidal and dissolved P 255


P (mg l�1) determined as described by Murphy & Riley

(1962), and C is a fitting parameter for the desorbable amount

of soil P (mg kg�1). The parameter C resembles the pool of

instantly labile P (Q0) used by Hartikainen (1991) in a linear

relationship between quantity and intensity.

Setting s0 ¼ 0 in Equation (2) allows us to calculate a P

concentration c00 in solution that is in equilibrium with P

originally sorbed by the soil:

c00 ¼C

k0 S0max � C

� � : ð3Þ

In addition, we determined the amount of added P sorbed

to dispersible particles for all samples from site A and topsoil

samples from site H. We ultra-centrifuged the supernatants

from the P sorption experiments of soil samples from site A

and from the Ap horizon of site H. The Langmuir isotherms

were also calculated for the ultra-centrifuged (dissolved) P

concentrations. Samples that were centrifuged at 3000 g are

denoted centrifuged and those at 300 000 g ultra-centrifuged.

Fitted parameters of the isotherms were analysed for correlation

with soil properties by the non-parametric Tau statistic of Kendall.

Dispersible P

To determine the concentration of dispersible P, 10 g of soil

was shaken end-over-end at 30 r.p.m. with 20ml 0.01M KCl

for 24 hours. After centrifuging at 3000 g for 10minutes, the

supernatant was filtered through P-free filter papers (Schleicher

and Schuell, No 512½). An aliquot of this filtrate was

ultra-centrifuged at 300 000 g for 1hour at 10�C (Beckman

Optima TL, Unterschleissheim, Germany). The concentration

of dispersible P was calculated as the difference between P

concentrations of ultra-centrifuged and centrifuged samples.

Results

Concentrations of dissolved P in drainage water and

groundwater

The variation in the concentration of soluble reactive P in

drainage water was large. Concentrations ranged from

<10�g l�1 to 1700�g l�1 with no evident seasonal trend.

Therefore, the data were pooled to give depth distributions

of concentrations displayed in Figure 1. Median concentra-

tions were <20�g l�1 for most sites and sampling depths.

Median concentrations >100�g l�1 were found down to

5.6m at site A, with an exception for samples from 1.5m

depth. Concentrations >100�g l�1 were also detected in sam-

ples from the lower part of the aquifer (Figure 2).

Concentrations of P in soil and degree of P saturation

The total P concentration of the arable soils A and H was

approximately 190mgkg�1 larger than the concentrations of the

fallow sites D and S at 0–30 cm (Table 2). In the 60–90 cm layer,

the total P concentration of site H was within the range of total P

A0

1

2

3

4

5

6

0 100 200 300 400

Dep

th b

elow

soi

l sur

face

/m

0

1

2

3

4

0 100 200 300 400

D S

Fallow sites

n n

57

59

59

12

13

11

4996

10

Hn n6771

5210

14

65

70

699

14

12

14

Arable sites

Concentration of soluble reactive P /µg l–1

Figure 1 Depth distribution of concentrations of

soluble reactive P in drainage water. The whiskers

of the box plots indicate 10th and 90th percentiles,

the boxes range from 25th to 75th percentiles. The

median splits each box in two parts. Data from 18

May 2000 to 4May 2001.. indicate the minimum

depth of groundwater tables, ! the minimum

depth of perched water tables.



concentrations of the fallow sites, whereas the concentration at

site A was more than twice the concentration of the other sites.

Differences between fertilized and fallow plots were less pro-

nounced for concentrations of oxalate-extractable P in 0–30 cm,

but clear for 30–60 cm depth. Again, the concentration of oxalate-

extractable P in the 60–90 cm layer of site A was more than twice

the concentration of the other sites. Very small concentrations of

oxalate-extractable P were found in subsoil of site S.

Concentrations of Feox and Alox generally decreased with

increasing depth and differed little between arable and fallow

sites. Instead, larger subsoil concentrations of oxalate-extractable

Fe at sitesH and S than at sitesA andD reflect differences in parent

material. Soils at sites H and S developed from glacial sands,

whereas the soils of sites A and D developed from aeolian sands.

The degree of P saturation (Z) was >0.25 for the topsoils of

all sites. It decreased sharply with increasing depth at sites D,

H and S, but only slightly at site A.

Sorption isotherms

With increasing depth, the affinity parameter (k0) increased,

whereas the parameter for the desorbable amount of P (C)

decreased (Figure 3, Table 3). Both parameters were signifi-

cantly correlated with the concentrations of oxalate-extractable

P and total P. Additionally, C was positively correlated with

the degree of P saturation (r¼ 0.67).

The maximum concentration of adsorbable P (S0max) was

not significantly correlated with the concentrations of either

total P or oxalate-extractable P, or with the degree of P satur-

ation. However, S0max of subsoil samples of site A, showing

exceptional large concentrations of soil P, was only 24–47% of

that of samples of the other sites (Table 3).

Calculated equilibrium concentrations decreased with increas-

ing depth (Table 4). For depths >30 cm, those of the arable sites

were much larger than those of fallow sites. Overall, they were

similar to the concentrations we determined by shaking the soil

with 0.01M KCl (Table 4). Furthermore, calculated equilibrium

concentrations and concentrations determined in dilute KCl for

soil samples were similar to the median soluble reactive P con-

centrations of drainage water from 90cm depth (Table 4).

Colloidal P in soils, drainage water, and groundwater

We could not fit the modified Langmuir sorption isotherm,

Equation (2), to the data that were obtained for centrifuged

soil samples from 0–30 cm depth at site A (Figure 3). Up to a

concentration of 30mg l�1 in solution, the addition of ortho-

phosphate caused a desorption of P. Ultra-centrifuging of the

ML 7

Concentration of soluble reactive P /µg l–1

0 100 200 300D

epth

/m

0

10

20

30

40

ML 8

0 100 200 300

Figure 2 Depth distribution of concentrations of

soluble reactive P in groundwater samples.

Samples were collected on 16 December 1999

(.), 20 May 2000 (s) and 4 May 2001 (.).

Table 2 Phosphorus, iron and aluminium concentrations of the soils

Total P Poxa Feox

a Aloxa

Site /mgkg�1 Zb

0–30 cm depth

A 804 (14) 597 (12) 1531 (79) 835 (13) 0.69

H 884 (3) 493 (62) 1830 (108) 824 (26) 0.50

D 693 (22) 530 (8) 1559 (55) 736 (9) 0.62

S 610 (99) 291 (32) 661 (79) 1453 (61) 0.29

30–60 cm depth

A 397 (2) 342 (26) 315 (12) 794 (43) 0.63

H 325 (1) 243 (24) 941 (94) 763 (23) 0.35

D 244 (19) 182 (5) 506 (118) 1105 (50) 0.24

S 84 (28) 7 (2) 785 (20) 550 (8) 0.01

60–90 cm depth

A 190 (55) 145 (16) 149 (3) 381 (13) 0.56

H 81 (54) 68 (4) 466 (26) 456 (25) 0.17

D 90 (36) 70 (14) 234 (44) 733 (14) 0.14

S 41 (11) ND 745 (60) 375 (9) –

aOxalate-extractable P, Fe and Al.bZ¼ [Pox]/0.5([Feox]þ [Alox]), with [Pox], [Feox] and [Alox] in mmol

kg�1 (van der Zee & van Riemsdijk, 1986).

ND, below detection limit.

The numbers in parentheses denote the difference between two analy-

tical replicates for total P and the standard error of four analytical

replicates for Pox, Feox and Alox.



batch solution samples from the Ap horizon of site A

decreased soluble reactive P concentrations in the supernatant

up to 95% (mean 75%, Figure 3). For the subsoil samples,

isotherms from ultra-centrifuged samples had larger S0max and

smaller affinity constants than isotherms for non-ultra-centrifuged

samples (Table 3). Differences between sorption isotherm

parameters from centrifuged and ultra-centrifuged samples

were much smaller for site H than for site A.

The absolute concentration of colloidal P increased with the

amount of P sorbed to the soil (Figure 4). A graph of colloidal

P concentrations against concentrations of oxalate-extractable

P suggests a threshold concentration of P in the soil for the

dispersion of colloidal P (Figure 5). The concentrations of

organic C in the soil, soil pH and the degree of P saturation,

Z, seemed to influence concentrations of colloidal P to a

smaller extent.

In general, ultra-centrifuging did not reduce concentrations of

soluble reactive P in drainage water significantly (Figure 6). Large

concentrations of colloidal P were found on 24 February 2001 and

26 February 2002 in drainage water from the sampler at 90 cm

depth at site A that collected the largest cumulative volume of

drainage water among replicates (replicate 5, Figure 6).

Discussion

Leaching of dissolved P and its controls

From an ecological point of view, the median concentrations

of P in drainage water from sites H, D and S are typical of

mesotrophic surface waters (OECD, 1982) and background

concentrations in the catchment of the river Spree (Driescher

& Gelbrecht, 1993). Median concentrations of soluble reactive

P of drainage water from site A are mostly larger than the

critical threshold of 100�g l�1 used in the Netherlands

(Breeuwsma et al., 1995). They are larger than the German

guideline value of 150�g l�1 orthophosphate in three of six

sampling depths (Auerswald et al., 2002).

Centrifuged

–50

0

50

100

150

Ultra-centrifuged

30–60 cm

Sor

bed

or d

esor

bed

P /m

g kg

–1

–50

0

50

100

150

Soluble reactive P /mg l–1

0 10 20 30 40 50–50

0

50

100

15060–90 cm

0 10 20 30 40 50

0–30 cm

A

H

D

S

Figure 3 Modified Langmuir sorption iso-

therms, Equation (2), for soil samples.

Concentrations greater than 50mg l�1 for

samples from site A are not shown so that

details of smaller P concentrations are clear.

Data points are the arithmetic mean of duplicate

determinations. Error bars denote the range

between duplicates. The difference between

duplicates is often smaller than the symbol size.



The depth distributions of P concentrations in groundwater

integrate over larger areas than data for the vadose zone. The

profiles we obtained resemble those in areas with little agri-

culture in the Spree catchment (Driescher & Gelbrecht, 1993).

Concentrations > 100�g l�1 that we found in groundwater as

well as those recorded by Driescher & Gelbrecht (1993) were

restricted to depths where anaerobic conditions prevail, as

indicated by the absence of nitrate (for depth profiles of nitrate

see Siemens et al., 2003). Driescher & Gelbrecht (1993)

explained this as resulting from the release of P by reductive

dissolution of iron oxides.

We found a good agreement of median concentrations of

soluble reactive P in drainage water, calculated equilibrium

concentrations of soil samples, and concentrations measured

in extracts with dilute KCl (Table 4). This indicates that solu-

ble reactive P of drainage water is in equilibrium with P sorbed

Table 3 Parameters of the fitted Langmuir isotherms, Equation (2)

With colloidal P (non-ultra-centrifuged) Without colloidal P (ultra-centrifuged)

C S0max k0 C S0

max k0

Site /mg kg�1 /mg kg�1 /l mg�1 R2 /mgkg�1 /mgkg�1 /l mg�1 R2

0–30 cm depth

A – – – – 113 205 0.70 0.89

H 23 147 0.10 0.98 26 225 0.12 0.94

D 39 120 0.20 0.99 – – – –

S 40� 10�9 96 1.70 0.75 – – – –

30–60 cm depth

A 9 55 0.57 0.94 15 787 0.05 0.80

H 7 117 0.28 0.99 – – – –

D 5 123 0.80 0.98 – – – –

S 2 132 2.97 0.99 – – – –

60–90 cm depth

A 7 32 2.38 0.80 4 162 0.33 0.85

H 4 133 1.09 0.99 – – – –

D 6� 10�9 112 1.03 0.97 – – – –

S 1 128 3.95 0.98 – – – –

All regressions were significant at P <0.05.

Table 4 Comparison of calculated equilibrium concentrations, c00 of Equation (3), concentrations of extracts with dilute solution of KCl, and median

soluble reactive P concentrations in drainage water

Site

Calculated equilibrium concentration

/�g l�1

Concentration measured

in 0.01M KCla /�g l�1

Median concentration

in drainage water /�g l�1

0–30 cm depth

A – 4120 (220) –

H 1854 1650 (300) –

D 2407 1970 (18) –

S 0 50 (10) –

30–60 cm depth

A 343 395 (50) –

H 227 250 (90) –

D 53 100 (30) –

S 5 10 (0) –

60–90 cm depth

A 118 100 (0) 104

H 28 40 (0) 19

D 0 30 (0) 37

S 2 10 (10) 7

aNumbers in parentheses denote the difference between duplicate determinations.



to the solid phase. McDowell & Sharpley (2001) found a

similar agreement of P concentrations in drainage from top-

soils and P concentrations in batch extracts with 0.01M CaCl2.

Accordingly, the degree of P saturation (Z) and S0max, which

rely on equilibrium, were good indicators of P leaching. They

differentiated clearly between sites D, H and S with median

concentrations <100�g l�1 and site A with median concentra-

tions >100�g l�1. Similarly, McDowell et al. (2002) showed

that a degree of P saturation (Q/Qmax)> 0.4 caused large con-

centrations of P in drainage from topsoil. Next to the degree of

P saturation, the Langmuir affinity constant (k) was a sensitive

indicator of large concentrations of P in drainage in the study

of McDowell et al. (2002). In contrast, the affinity constant (k0)

of our modified Langmuir equation (2) was not related to

concentrations of soluble reactive P in drainage water. The

constant derived from soil samples at 60–90 cm at site A was

Sorbed P (ultra-centrifuged) /mg kg–1

0 20 40 60 80 100 120 140

Col

loid

al P

in s

uper

nata

nt /m

g l–1

0

10

20

30

40

50

60

A 0–30 cm

A 30–60 cm

A 60–90 cm

H 0–30 cm

Figure 4 Concentrations of colloidal P in the

supernatant of batch extracts as a function of

the change in the total concentration of sorbed

P (sorbed to colloids and to the non-suspended

solid phase). Data points are the arithmetic

means of duplicate determinations. Error bars

denote the difference between duplicates.

Z

Z

0.25 0.50 0.75

0

1000

2000

3000

Corg

Organic C /%

0 0.5 1.0 1.5

pH

pH

4.0 4.5 5.0 5.5

Col

loid

al P

con

cent

ratio

n in

sup

erna

tant

/µg

l–1

0

1000

2000

3000

A

H

D

S

Pox

Oxalate-extr. P /mg kg–1

0 250 500 7500

Figure 5 Influence of different soil properties

on the concentration of colloidal P released in a

0.01M KCl solution. Data points are the

arithmetic means of duplicate determinations.

Error bars denote the difference between

duplicates.



larger than the constants of sites H and D. By normalizing

sorption to S0max (s

0/S0max), we derived an exponential relation-

ship between P concentration in drainage water and the

affinity parameter k0 similar to the one of McDowell et al.

(2002). In our case, however, concentrations of soluble reactive

P increased greatly if k0 < 0.7 lmg�1; this critical value is an

order of magnitude larger than the value of 0.07 lmg�1 found

by McDowell et al. (2002) for k.

In contrast to our results for the sandy soils from Munster,

concentrations of soluble reactive P of subsoil drainage were

not controlled by sorption equilibrium in the loamy soils

investigated by Heckrath et al. (1995) and Anderson & Xia

(2001). Although the subsoils in their studies were not satur-

ated with P and showed no alteration of sorption character-

istics upon fertilization, they found that applying P affected

concentrations of P in the drainage water. They explained the

effect by a lack of interaction between solute and solid phase

because of preferential flow. In Munster, preferential flow

might be one reason for the observed variation of soluble

reactive P concentrations because we know that the suction

plates sample preferential flow (our own unpublished data).

Preferential flow might be responsible for large concentrations

of P at 5.6m at site A. However, samples from depths >2.2m

were collected intermittently in suction cups, which are

unlikely to sample preferential flow. Another factor that prob-

ably increased the variationofP concentrationsbetween samplers

at any one sampling depth as well as between different depths

at one site is the heterogeneity of the soil. The greater the

depth of sampling, the more uncertain is the spatial origin of

the collected drainage water. This is especially true for sites A

and D, where perched water tables might easily cause lateral

transport.

Mobilization and transport of colloidal P

Especially at large degrees of P saturation, a large fraction of

added P was sorbed to dispersed particles (Figure 3), which

may be related to the large surface area of colloids (Kretzschmar

et al., 1999). Our results imply that adsorption of P

increases the dispersibility of P bound to particles (Figures 4

and 5). This accords with the shift of the surface potential of

iron oxides, clay minerals, and calcite to negative values

caused by the adsorption of orthophosphate (Stumm & Sigg,

1979; Puls & Powell, 1992; Celi et al., 1999, 2000). Given a soil

pH of 5.5–6.4 (Table 1), the sorption of P might, for example,

induce a reduction and ultimately a reversal of the surface

charge of goethite from positive to negative (Stumm & Sigg,

1979). Goethite particles associated with negatively charged

clay minerals or organic matter as a result of electrostatic

forces might in turn become detached and mobilized.

The mobilization of colloidal P by addition of P in fertilizer

can have important environmental consequences. Additions of

P to soils in excess of crop requirements can saturate the soils’

sorption capacity for P and can additionally mobilize P that

Soluble reactive P concentration before ultra-centrifuging /µg l–1

0 100 200 300 400 500

Con

cent

ratio

n af

ter

ultr

a-ce

ntrif

ugin

g /µ

g l–

1

0

100

200

300

400

500

A

H

D

S

ML wells

A, 90 cm depthreplicate 5

26 February 2002

24 February 2001

1:1 line

n = 136

Figure 6 Soluble reactive P concentrations of

drainage water and groundwater before and

after ultra-centrifuging.



was sorbed in the past. In other words, over-fertilization with

P may cause a part of the P retained in a stationary phase in

the soil to become mobile.

Using KCl as background electrolyte we determined the

mobilization potential of colloids. Large concentrations of

Ca2þ that dominate over concentrations of monovalent

cations might reduce the actual mobilization and mobility of

colloids under field conditions. In general, samples of drainage

water contained only small concentrations of colloidal P

(Figure 6). Leaching of colloidal P seems to be restricted to

channels carrying large volumes of drainage water (sampler

replicate 5). Furthermore, it seems to have been temporally

restricted to February, when sampler replicate 5 collected

approximately 37% of its cumulative sample volume of the

winter. These results accord with the findings of Kaplan et al.

(1993) that concentrations of mobile colloids increase and the

colloid diameter decreases with increasing flow rate.

Because nitrate had been flushed from the soil profile, total

electrolyte concentrations at 90 cm depth at site A were on

average only 104�S cm�1 in February compared with a grand

mean of 230�S cm�1 (for time series of nitrate concentrations

see Siemens & Kaupenjohann, 2002). The electrical conductiv-

ity of the two samples of sampler replicate 5 at 90 cm depth

that contained large concentrations of colloidal P was 166�S

cm�1 (24 February 2001), and 86�S cm�1 (26 February 2002).

These conductivities are in the range 30–170�S cm�1 within

which Kaplan et al. (1996) found dispersed particles.

P budgets of subsoils

If we take Z¼ 0.01 as the background degree of saturation, as

at site S, then we can calculate a background P concentration of

4mgkg�1 at site A, 7mgkg�1 for site D, and 5mgkg�1 for site

H. It then follows that the actual concentrations of oxalate-

extractable P are the result of an accumulation of 141mgP kg�1

(site A) and 63mgP kg�1 (sites D and H) in the subsoil horizons

below the plaggic horizons. The groundwater recharge at our

research sites is approximately 300mmyear�1 (Siemens et al.,

2003). If the concentration of P were 350�g l�1 (c00 A, 60 cm

depth) and the soils’ bulk density was 1.4kg l�1 then it would

take 376 years to achieve the accumulation of P in 70–90 cm at

site A. For the sites H (60–90 cm, c00 ¼ 240�g l�1) and D

(70–90 cm, c00 ¼ 55�g l�1), these times would be 368 years and

1069 years, respectively. Despite the former addition of Plaggen

material, these times seem overly long because we extrapolated

the large present concentrations of soluble reactive P to the past.

It is therefore likely that other forms of P, which might be

colloidal P or dissolved organically bound P, were leached

from the soils together with soluble reactive P.

Conclusions

Concentrations of soluble reactive P in excess of 100�g l�1

down to depths of 33m show that P is transported locally and

lost from the soil in the catchment we studied.

The median concentration of soluble reactive P in drainage

water is controlled by rapid adsorption in the sandy soils.

Therefore, equilibrium concentrations of P determined in

batch experiments and a threshold of the degree of P satura-

tion of 0.25 will be good predictors of P concentrations of

drainage water.

Accumulation of P in soils increases the risk that P sorbed

on colloids will be leached when large volumes of water pass

through the soil, especially if the total concentration of

electrolyte is small.

Acknowledgements

We thank Anke Schwolow, Nadine Kurowski, Kristine

Schimpff, Christian Mertens, Martti Haas, Andre Meller and

Alfons Peine for their assistance. We are grateful to Christoph

and Heinz Ahlert, Andreas and Egon Henrichmann, Georg

Schulze-Dieckhoff and the Stadtwerke Munster GmbH for

permission to do the experiments on their properties. This

study was financed by the agriculture and water works

cooperation programmes of the City of Munster and the City

of Warendorf, the Stadtwerke Munster GmbH, the

Wasserversorgung Beckum GmbH, the Westfalisch-Lippischer

Landwirtschaftsverband and the Environmental Department

of the City of Munster.

References

Anderson, R. & Xia, L. 2001. Agronomic measures of P, Q/I param-

eters and lysimeter collectable P in subsurface soil horizons of a

long-term slurry experiment. Chemosphere, 42, 171–178.

Auerswald, K., Claasen, N., Romer, W. & Werner, W. 2002.

VDLUFA-Standpunkt: Mogliche Folgen hoher Phosphatgehalte

im Boden und Wege zu ihrer Verminderung. Mitteilungen der

Deutschen Bodenkundlichen Gesellschaft, 98, 75–80.

Beek, J. 1978. Phosphate retention by soil in relation to

waste disposal. Doctoral dissertation, Agricultural University,

Wageningen.

Bottcher, A.B., Miller, L.W. & Campbell, K.L. 1984. Phosphorus

adsorption in various soil water extraction cup materials: effect of

acid wash. Soil Science, 137, 239–244.

Breeuwsma, A. & Silva, S. 1992. Phosphorus Fertilisation and Environ-

mental Effects in the Netherlands and the Po Region (Italy). Report

57, DLO The Winand Staring Centre for Integrated Land, Soil and

Water Research, Wageningen.

Breeuwsma, A., Reijerink, J.G.A. & Schoumans, O.F. 1995. Impact of

manure on accumulation and leaching of phosphate in areas of

intensive livestock farming. In: Animal Waste and the Land–Water

Interface (ed. K. Steele), pp. 239–249. CRC/Lewis Publishers, Boca

Raton, FL.

Celi, L., Lamacchia, S., Marsan, F.A. & Barberis, E. 1999. Interaction

of inositol hexaphosphate on clays: adsorption and charging

phenomena. Soil Science, 164, 574–585.

Celi, L., Lamacchia, S. & Barberis, E. 2000. Interaction of inositol

phosphate with calcite. Nutrient Cycling in Agroecosystems, 57,

271–277.



Driescher, E. & Gelbrecht, J. 1993. Assessing the diffuse phosphorus

input from subsurface to surface waters in the catchment area of the

lower River Spree (Germany). Water Science and Technology, 28,

337–347.

Hartikainen, H. 1991. Potential mobility of accumulated phosphorus

in soil as estimated by the indices of Q/I plots and by extractant.

Soil Science, 152, 204–209.

Haygarth, P.M., Warwick, M.S. & House, W.A. 1997. Size distribu-

tion of colloidal molybdate reactive phosphorus in river waters and

soil solution. Water Research, 31, 439–448.

Heckrath, G., Brookes, P.C., Poulton, P.R. & Goulding, K.W.T. 1995.

Phosphorus leaching from soils containing different phosphorus

concentrations in the Broadbalk experiment. Journal of Environmental

Quality, 24, 904–910.

Hens, M. & Merckx, R. 2001. Functional characterization of colloidal

phosphorus species in the soil solution of sandy soils. Environmental

Science and Technology, 35, 493–500.

Kaplan, D.I., Bertsch, P.M., Adriano, D.M. & Miller, W.P. 1993. Soil--

borne mobile colloids as influenced by water flow and organic carbon.

Environmental Science and Technology, 27, 1193–1200.

Kaplan, D.I., Sumner, M.E., Bertsch, P.M. & Adriano, D.C. 1996.

Chemical conditions conducive to the release of mobile colloids from

Ultisol profiles. Soil Science Society of America Journal, 60, 269–274.

Kretzschmar, R., Borkovec, M., Grolimund, D. & Elimelech, M. 1999.

Mobile subsurface colloids and their role in contaminant transport.

Advances in Agronomy, 66, 121–193.

Laubel, A., Jacobsen, O.H., Kronvang, B., Grant, R. & Andersen, H.E.

1999. Subsurface drainage loss of particles and phosphorus from

field plot experiments and a tile-drained catchment. Journal of

Environmental Quality, 28, 576–584.

Leinweber, P., Lunsmann, F. & Eckhardt, K.U. 1997. Phosphorus

sorption capacities and saturation of soils in two regions with

different livestock densities in northwest Germany. Soil Use and

Management, 13, 82–89.

McDowell, R.W. & Sharpley, A.N. 2001. Approximating phosphorus

release from soils to surface runoff and subsurface drainage. Journal

of Environmental Quality, 30, 508–520.

McDowell, R., Sharpley, A. & Withers, P. 2002. Indicator to predict

the movement of phosphorus from soil to subsurface flow. Environ-

mental Science and Technology, 36, 1505–1509.

Miedema, R. 1991. Anthrosols. In: The Major Soils of the World (eds

P.M. Driessen & R. Dudal), pp. 44–50. Agricultural University,

Wageningen.

Murphy, J. & Riley, J.P. 1962. A modified single solution method for

the determination of phosphate in natural waters. Analytica

Chimica Acta, 27, 31–36.

OECD 1982. Eutrophication of Waters, Monitoring, Assessment and

Control. Organisation for Economic Co-operation and Develop-

ment, Paris.

Puls, R.W. & Powell, R.M. 1992. Transport of inorganic colloids

through natural aquifer material: implications for contaminant

transport. Environmental Science and Technology, 26, 614–621.

Schlichting, E., Blume, H.P. & Stahr, K. 1995. Bodenkundliches Prak-

tikum, 2nd edn. Blackwell, Berlin.

Sharpley, A.N., Chapra, S.C., Wedepohl, R., Sims, J.T., Daniel, T.C. &

Reddy, K.R. 1994. Managing agricultural phosphorus for protection

of surface waters: issues and options. Journal of Environmental

Quality, 23, 437–451.

Siemens, J. & Kaupenjohann, M. 2002. Contribution of dissolved

organic nitrogen to N leaching from four German agricultural

soils. Journal of Plant Nutrition and Soil Science, 165, 675–681.

Siemens, J. & Kaupenjohann, M. 2003. Dissolved organic carbon is

released from sealings and glues of pore water samplers. Soil Science

Society of America Journal, 67, 795–797.

Siemens, J., Haas, M. & Kaupenjohann, M. 2003. Dissolved organic

matter induced denitrification in subsoils and aquifers? Geoderma,

113, 253–271.

Staude, H. 1986. Erlauterungen zu Blatt 3911 Greven – Geologische Karte

Nordrhein-Westfalen 1:25000. Geologisches Landesamt, Krefeld.

Stumm, W. & Sigg, L. 1979. Kolloidchemische Grundlagen der

Phosphor-Elemination in Fallung, Flockung und Filtration.

Zeitschrift fur Wasser- und Abwasser-Forschung, 2/79, 73–83.

Van der Zee, S.E.A.T.M. & van Riemsdijk, W.H. 1986. Transport of

phosphate in a heterogeneous field. Transport in Porous Media, 1,

339–359.

Van der Zee, S.E.A.T.M. & van Riemsdijk, W.H. 1988. Model for

long-term phosphate reaction kinetics in soil. Journal of Environ-

mental Quality, 17, 35–41.

Van der Zee, S.E.A.T.M., Nederlof, M.M., van Riemsdijk, W.H. & de

Haan, F.A.M. 1988. Spatial variability of phosphate adsorption

parameters. Journal of Environmental Quality, 17, 682–688.

Werner, W. 1997. Implementation and efficiency of countermeasures

against diffuse nitrogen and phosphorus input into groundwater

and surface waters. In: Controlling Mineral Emissions in European

Agriculture (eds E. Romstad, J. Simonsen & A. Vatn), pp. 73–88.

CAB International, Wallingford.



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