Pedosphere 27(1): 1–16, 2017

doi:10.1016/S1002-0160(15)60106-0

ISSN 1002-0160/CN 32-1315/P

c⃝ 2017 Soil Science Society of China

Published by Elsevier B.V. and Science Press

Biosolid Application to Agricultural Land—a Contribution toGlobal Phosphorus Recycle: A Review

Silvana Irene TORRI1,∗, Rodrigo Studart CORREA2 and Giancarlo RENELLA3

1Department of Natural Resources and Environment, School of Agriculture, University of Buenos Aires, Avenue San Martin 4453,

Buenos Aires 1417 DSE (Argentina)2University of Brasılia-UnB/PPGCA, Campus Darcy Ribeiro, Caixa Postal 04.401, 70910-970 DF (Brazil)3Department of Agrifood Production and Environmental Sciences, University of Florence, Piazzale delle Cascine 18, Florence 50144

(Italy)

(Received June 30, 2016; revised November 10, 2016)

ABSTRACT

Phosphorus (P) is an essential nutrient required for plant development. Continuous population growth and rising global demand

for food are expected to increase the demand for phosphate fertilizers. However, high-quality phosphate rock reserves are progressively

becoming scarce. Part of the increased pressure on P resources could be alleviated by recycling P present in biosolids. Therefore, it is

crucial to understand the dynamics of P in biosolid-amended soils, the effects of residual biosolid-borne P in soils, the way in which

microorganisms may control P dynamics in biosolid-amended soils and the environmental implications of the use of biosolids as a

source of P. Further research is needed to maximize biosolid-borne P uptake by crops and minimize its loss from biosolid-amended

soils. The analysis of the microbiological control of P dynamics in biosolid-amended soils indicates interactions of biosolid P with

other nutrients such as carbon (C) and nitrogen (N), suggesting that harmonization of the current regulation on the use of biosolids

in agriculture, mainly based on total N and pollutant contents, is needed to better recycle P in agriculture.

Key Words: anthropogenic P, phosphate, P availability, P biogeocycle, P uptake, runoff P

Citation: Torri S I, Correa R S, Renella G. 2017. Biosolid application to agricultural land—a contribution to global phosphorus

recycle: A review. Pedosphere. 27(1): 1–16.

INTRODUCTION

Phosphorus (P) is an essential nutrient for all forms

of life. Biomolecules containing P are present in cellu-

lar components, including membranes (phospholipids),

genetic material (DNA and RNA), and energy storage

(ATP and ADP), among others (Elser, 2012). While

humans and animals satisfy their need for P via food

intake, plants have to absorb it from soils. In spite of

its wide distribution in nature, P is one of the least

available mineral nutrients to plants (Goldstein et al.,

1988), and P uptake is usually a growth-limiting fac-

tor (Grant et al., 2005). Unlike nitrogen (N), the bio-

geochemical cycle of P does not include a significant

gaseous component, since its annual atmospheric de-

position rates are in the order of 0.25 kg P ha−1 year−1

(Liu Y et al., 2008). In natural ecosystems, P is entire-

ly supplied from the weathering of parent materials

(Schlesinger and Bernhardt, 1997), and the amount

of total P is preserved because it is released back to

the soil system through plant residues, animal excreta

or when organisms die. In agricultural systems, crop

removal represents the primary route by which P is

lost from soils. Unless P sources are artificially incor-

porated to agricultural soils, both total and available

P stocks steadily decrease with time to the point that

the soil can no longer adequately supply plant P needs

(Van Vuuren et al., 2010). In the course of time, soil P

depletion may lead to loss of soil fertility and produc-

tivity.

Mineral phosphate fertilizers are the primary so-

urce of P input to agricultural lands. Even though the

use of rock phosphate-based fertilizers was introduced

in the 1820s, it was not until the late 1940s that P

fertilizers were increasingly requested. In 2011, global

phosphate fertilizer production resulted in the deple-

tion of approximately 20 Mt of P from phosphate rock

(Jasinski, 2013). The demand for P is expected to in-

crease in the following years due to continuous popu-

lation growth and rising global demand for food, with

a predicted increase to approximately 257 Mt by 2017

(Heffer, 2013; Jasinski, 2013). Economic, high-quality

∗Corresponding author. E-mail: torri@agro.uba.ar.

2 S. I. TORRI et al.

phosphate rock reserves are progressively becoming

scarce (Cordell and Neset, 2014). Although reserves of

phosphate rock are found in several countries and new

reserves have been identified (Midgley, 2012), phos-

phate rock is a finite, non-renewable resource. Accor-

ding to the U.S. Geological Survey, phosphate deposits

will last about 50 years at the current rate of extrac-

tion (Kelly and Matos, 2013). Therefore, there is an

increasing concern regarding phosphate rock reserves

to become depleted.

Part of the increased pressure on P resources could

be alleviated by recycling P present in various agricul-

tural and urban wastes (Frossard et al., 2009; Mac-

Donald et al., 2011). However, the joint effects of poor

knowledge of P status and the lack of a clear regulation

on manure or organic waste agricultural management

still limits P recycling potential in agriculture. This

paper reviews the availability and environmental fate

of P present in biosolids and envisages some possible

strategies for its sustainable management.

BIOSOLIDS AS A SOURCE OF P

Land application of organic by-products is an eco-

nomically attractive waste management strategy, lar-

gely promoted by scientists and regulating organisms.

Furthermore, it has been a socially accepted practice

for decades in many parts of the world (Tsadilas, 2011;

Larney and Angers, 2012; Lu et al., 2012).

The term biosolid was introduced in the early 1990s

to designate the solid, semi-solid or liquid materials

generated from the treatment of domestic sewage slu-

dge that has been sufficiently processed to be safely

land-applied. Biosolids contain organic carbon (C), N,

P, potassium (K), sulphur (S), calcium (Ca), magne-

sium (Mg), and microelements necessary for plants and

soil fauna to live. Nutrient contents in biosolids depend

on the untreated water source, chemicals used for pu-

rification, and types of unit operations used, and were

reported to be in the ranges of 1–210 g N kg−1, 1–

150 g P kg−1, 1–65 g K kg−1, 5–170 g Ca kg−1, and

2–94.5 g Mg kg−1 (Hansen and Chaney, 1984; Solis-

Mejia et al., 2012). Application of biosolids on agricul-

tural and degraded lands is one of the most promising

alternatives of disposal, because it offers the possibility

of recycling plant nutrients and organic matter (Gar-

cıa-Orenes et al., 2005; Torri and Lavado, 2009a, b;

Kowaljow et al., 2010). This practise may also con-

tribute to soil C sequestration, reducing greenhouse gas

emissions (Haynes et al., 2009; Tian et al., 2009; Torri

and Lavado, 2011; Torri et al., 2014). However, bio-

solids may contain undesirable hazardous substances

such as potentially toxic trace elements ranging from

less than 1 to over 1 000 mg kg−1, polychlorinated

biphenyls (PCBs), polycyclic aromatic hydrocarbons

(PAHs), and dioxins (Abad et al., 2005; Martınez et

al., 2007; Torri, 2009; Ahumada et al., 2014; Jordan et

al., 2016). Consequently, biosolids have to be properly

treated and disposed to prevent health risk and en-

vironmental contamination (Kroiss, 2004). Although

to date experimental results indicate a low level of

risk for crops or pastures (Torri and Lavado, 2009a,

b; Cogger et al., 2013a), application of biosolids onto

non-agricultural land is usually preferred to avoid the

risk of hazardous substances entering the food chain

(Magesan and Wang, 2003; Athamenh et al., 2015).

In the European Union, a global regulation on

biosolid use in agriculture relies on the Water Frame-

work Directive (2000/60/EC) (EC, 2000) and the sub-

sequent Groundwater Directive (2006/118/EC) (EC,

2006), which have resumed all the previous specific

Directives on bathing waters, sewages sludge, urban

wastes and nitrates, and limit the potential recycle of

any biosolids in agriculture to their impacts on surface

water, groundwater, and atmosphere caused by exces-

sive nutrient, organic and inorganic pollutants. While

most organic pollutants can be degraded and excessive

N may be volatilized during sludge treatment, trace

elements are generally concentrated and may exceed

the mandatory limits for sludge application to agri-

cultural soils (CEC, 1986). Elevated contents of trace

elements prevent the use of sludge as a soil amendment

because of their negative impacts on soil microbial di-

versity and microbial activity (Renella et al., 2007a;

Gomes et al., 2010).

In wastewater, P is mainly found as orthophos-

phates, usually linked to small amounts of organic P

(Tran et al., 2012). Phosphorus removal is performed

by biological treatment or physiochemical precipita-

tion. In both cases, the soluble forms of P are con-

verted into a solid fraction, which can be an insolu-

ble salt or microbial biomass (De-Bashan and Bashan,

2004). Physiochemical precipitation removes dissolved

P phosphates by the addition of aluminium (Al),

iron (Fe), or calcium (Ca) compounds (Lee and Lin,

2007). The reaction is probably a combination of sur-

face adsorption onto metal hydroxides with chemi-

cal precipitation of the metal phosphate, producing

low P concentrations in the liquid phase (Elliott and

O’Connor, 2007). Biological P removal (BPR) pro-

cess relies on the use of a specific group of bacteria

that take up P in excess for their growth requirements

(Chen et al., 2013; Keating et al., 2016). The excess of

P is stored as intracellular granules of polyphosphate

(Grady et al., 2011), concentrating diluted P in waste-

BIOSOLID APPLICATION AND GLOBAL P RECYCLE 3

water by 10–50 times in bacterial aggregates (Yuan et

al., 2012).

Depending on the pre- or post-treatment used,

mean total P in biosolids was reported to be in the

range of 3.7–72.6 g P kg−1 on a dry weight base (Bar-

barick and Ippolito, 2007; Cordell, 2010). The addition

of some type of liming agent to stabilize biosolids may

result in lower total P (Christie et al., 2001). Since

more stringent N and P discharge limits have been im-

plemented on wastewater treatment plants (WWTP)

in environmentally sensitive areas, total P in biosolids

is expected to increase from current values (Clark et

al., 2010; Qin et al., 2015).

In biosolids, P exists in both soluble and insolu-

ble organic and inorganic P compounds (Tian et al.,

2012). Inorganic P is the predominant form, represen-

ting 70%–90% of total P (O’Connor et al., 2004; He

et al., 2010). Most P in biosolids is commonly in the

forms of aluminium phosphate (Shannon and Verghese,

1976), adsorbed onto ferric hydroxo-phosphate surfaces

(Jenkins et al., 1971), and hydroxyapatite or tricalcium

phosphate (Stumm and Morgan, 1970), with relatively

low water-soluble P as compared to total P (Brandt

et al., 2004). Organic P is mainly found as orthophos-

phate monoesters, orthophosphate diesters, phospho-

nates, phytates, and phospholipids (Hinedi et al., 1989;

He et al., 2010; Torri and Alberti, 2012).

It is well known that availability of P to plants

depends on the replenishment of labile P in the soil

solution from diverse soil fractions (Beck and Sanchez,

1994). In a general sense, P availability is defined as

those P compounds that are present in the correct

chemical forms to be taken up by plants during their

life cycle or taken up and used by living biological

organisms. The most significant P compound in terms

of availability is the orthophosphate anion, which is

associated with readily accessible short-term availabi-

lity for plants (Montalvo et al., 2015; Anand et al.,

2016). Phosphorus precipitation and dissolution reac-

tions greatly influence its concentration in the soil so-

lution, whereas organic P has to be hydrolyzed and

mineralized by microbial biomass to release orthophos-

phate anions. Hence, the importance of insoluble P

compounds rests entirely on their ability to buffer P

solution concentration or to become soluble in the soil

environment (McLaughlin, 1984).

Phosphorus availability in biosolids is strongly in-

fluenced by the wastewater treatment (WWT) proces-

ses (O’Connor et al., 2004; Elliott et al., 2005; White

et al., 2010). Sludge treatment with high Al and/or Fe

doses results in biosolids having low available P con-

centrations, with Fe and Al phosphates as dominant

P forms (Shober and Sims, 2007). Taking into account

that the solubility kinetics of these phosphate mine-

rals is extremely slow, it is unlikely that such mine-

rals, once formed, would readily release P into the

soil solution (Strawn et al., 2015). In fact, P in bio-

solids treated with Al and Fe was found to be less

soluble than P in untreated biosolids or commercial

fertilizers (Kyle and McClintock, 1995). Addition of

lime was reported to increase biosolid pH and decrease

the solubility of P by the formation of recalcitrant Ca-

phosphate minerals (Maguire et al., 2006; Shober et

al., 2006; Islas-Espinoza et al., 2014). Conversely, bio-

solids obtained by BPR exhibit both elevated total P

and water-extractable P when exposed to anaerobic

conditions (Stratful et al., 1999; Penn and Sim, 2002;

Ebeling et al., 2003). As the latter achieves effluent P

standards without the use of metal salts, the resultant

biosolids are typically low in Al and Fe contents and

their water-soluble P is higher than that of the other

treatments (Penn and Sims 2002; Brandt et al., 2004).

Heat-dried biosolids (non-BPR) were reported to

have the lowest P availability of all WWT processes.

Heat drying has been seen to reduce P extractability

by an average of 75% compared to dewatered proces-

ses (Smith et al., 2002a). Sarkar and O’Connor (2004)

found that heat-dried biosolids containing high levels

of Al and Fe have less than 10% water-soluble P. Ot-

her researchers reported that water-soluble P in heat-

dried biosolids was relatively low, in the range of 0.2%–

38%, as compared to total P (Frossard et al., 1996a;

Brandt et al., 2004). Smith et al. (2002b) indicated

that heat drying changes available P forms into low

soluble crystalline P minerals such as hydroxyapatite

and iron pyrophosphate. It was hypothesized that

the relatively low P bioavailability may be partly

attributed to slow physical breakdown of the pellets

(O’Connor and Sarkar, 1999; Smith et al., 2002b). In

the light of all these, O’Connor et al. (2004) suggested

grouping biosolids into three categories according to

biosolid-borne P availability relative to the inorganic

fertilizer triple superphosphate (TSP): low (< 25% of

TSP), moderate (25%–75% of TSP), and high (> 75%

of TSP). Their study identified biosolids produced with

conventional WWT processes as being in the moderate

category, BPR biosolids as being in the high category,

and biosolids with high total Fe and Al as being in the

low category. Current potential techniques for direct

speciation of P in soil and organic matrices have been

reviewed by Kruse et al. (2015).

DYNAMICS OF P IN BIOSOLID-AMENDED SOILS

The dynamics of biosolid-borne P differs from that

4 S. I. TORRI et al.

of mineral fertilizer P because not all the P in bio-

solids is phytoavailable in soil (Penn and Sims, 2002;

Codling, 2014). As mentioned above, biosolid P phy-

toavailability is closely related to its chemical forms in

the solid phase (Akhtar et al., 2005), which depends on

the composition of the wastewater entering the treat-

ment plant and the type of treatment process used (Sa-

blayrolles et al., 2010). Biosolid-borne P is often less

soluble and less plant available compared to soluble

phosphate fertilizer P (Tian et al., 2009). Jenkins et

al. (2000) estimated that almost 50% of phosphate in

most biosolid products is available for plant uptake du-

ring the first year.

When biosolids are land applied, different processes

occur, such as P sorption/desorption, microbial decom-

position of organic P, and dissolution/precipitation of

mineral P phases. Thus, changes in the forms and con-

centrations of biosolid-borne P occur upon biosolid in-

corporation. Depending on the biological and physico-

chemical properties of a soil, the rate of one of these

processes may be higher than the other one.

Several studies have reported an increase in bio-

available P levels in biosolid-amended soils in line with

biosolid application rates (Akhtar et al., 2012; Alleoni

et al., 2012; Hosseinpur and Pashamokhtari, 2013; Sha-

heen and Tsadilas, 2013). This increase may be at-

tributed to the high concentration of inorganic P in

the biosolids. However, these differences seem to be

less pronounced in P-enriched soils or soils that have

high affinity to retain P, such as those derived from cal-

careous parent material (Sarkar and O’Connor, 2004;

Ippolito et al., 2007). Other researchers reported that

the shift of biosolid-borne P from less labile to more

labile forms may also contribute to increases in soil P

availability to plants (Sui et al., 1999; Haney et al.,

2015).

Calcium-dominated biosolids result in higher con-

centrations of water-soluble P in biosolid-amended

soils (Brandt et al., 2004). Moreover, alkaline-stabi-

lized biosolids exhibit mean percentages of water-

extractable P statistically higher than conventionally

stabilized biosolids, which may be attributed to the

mineralization of organic P in biosolids. Many studies

have found that lime amendments increase soil orga-

nic C mineralization (Wong and Su, 1997; Torri et al.,

2003). Jokinen (1990) studied the influence of treat-

ment process on available soil P in biosolid-amended

soils and concluded that Al treatment reduces soil P

availability, whereas Ca treatment increases soil P ava-

ilability. Recently, Withers et al. (2015) observed that

Fe-treated and thermally dried biosolids give the lowest

increases (3%–6%), whereas lime-treated biosolids pro-

duce the largest increases in available P (11%–12%).

Past research has shown that P availability in soils

amended with chemically treated, anaerobically diges-

ted biosolids followed the order: Ca treatment > Fe

treatment > Al treatment (Soon and Bates, 1982). An

explanation to this is that P is bound to soil Ca sur-

faces with lower binding energy compared with Fe or

Al surfaces which bind P more strongly (Delgado and

Torrent, 1997; Elliot et al., 2002b). Therefore, chemi-

cal addition of Al, Fe, and Ca during wastewater and

biosolid processing is of main importance in determi-

ning P dynamics in biosolid-amended soils. Moreover,

if biosolids had undergone thermal treatment, the re-

activity of Fe-bound P minerals in biosolids is consi-

derably reduced and, consequently, the release of avai-

lable P is restricted (Hogan et al., 2001).

Biosolid soil application may also enhance P mi-

neralization, which would contribute to releases of or-

ganic biosolid-borne P and thus lead to an increase

in extractable P levels in the soil (O’Connor et al.,

2004). The positive correlations between soil respira-

tion and labile organic C and N in biosolid-amended

soils suggest that the stimulation of the activity of both

soil and biosolid-borne microbial communities can e-

xert a general solubilizing effect towards biosolid nutri-

ents (Sanchez-Monedero et al., 2004; Jin et al., 2011),

including P (Haney et al., 2015). However, the inte-

raction between biosolid degradation and P availability

may be more complex. For instance, in a leaching ex-

periment, Silveira and O’Connor (2013) reported that

dissolved organic C (DOC) released through biosolid

mineralization does not follow the same pattern as P

in leachates. Their results suggest that part of the mi-

neralized P is sorbed onto Al and Fe oxides present in

the soil, masking any relationship between DOC and

water-extractable P concentrations.

Land application of biosolids also incorporates non-

crystalline, colloidal amorphous forms of Fe and Al

oxides with a large specific surface area (Shober et

al., 2006). These amorphous complexes of Fe and Al

oxides can effectively adsorb or bind native soil phos-

phates (Nanzyo, 1986). Most research results have

shown that land application of biosolids modifies not

only soil P adsorption capacity (Maguire et al., 2000;

Lu and O’Connor, 2001), but also certain soil proper-

ties, such as dissolved organic matter, electrical con-

ductivity, pH and biological properties (Silveira et al.,

2003; Gilmour et al., 2003; Torri, 2009; Scharenbroch

et al., 2013). Biosolid organic matter was found to have

an indirect effect on phosphate adsorption, through in-

BIOSOLID APPLICATION AND GLOBAL P RECYCLE 5

hibiting Al oxide crystallization and even through in-

creasing the amorphous nature of Al (Bøen et al.,

2013; Maguire et al., 2000). This effect was similar,

but less pronounced, for Fe compounds (Borggaard

et al., 1990). In addition, P availability in biosolid-

amended soils may be also modified by environmental

conditions, such as temperature and moisture content.

Specific reactions between biosolid-borne P and the

soil matrix increase with time and, thus, P extractabi-

lity may be considerably reduced. For example, Silve-

ira and O’Connor (2013) observed that P becomes

less bioavailable with time due to increased P sorp-

tion. Consequently, it is very complicated to predict

the dynamics and availability of biosolid-borne P in

biosolid-amended soils.

EFFECT OF RESIDUAL BIOSOLID P IN SOILS

Phosphorus from biosolids has been applied in ex-

cess to soils at N-based rates because sewage materials

have a much higher P:N ratio (0.5–1.1) than required

by plants (0.07–0.14) (Mitchell et al., 2000). Even

though the fate of P after biosolid N exhaustion is

still an unsolved matter in the management of sewage

materials in soils (Correa, 2004). Nutrients from bio-

solids and chemical fertilizers may continue to act in

soils beyond the period they are supposed to promote

plant growth (Weatherley et al., 1988). Such a linge-

ring effect to nourish plants decreases with time (Cor-

rea and Silva, 2016) to a negligible level considered

as residual effect (Barrow and Campbell, 1972). Seve-

ral studies have measured residual effects of fertilizers,

but only few have done it for biosolids (Michael et al.,

1991). Residual effects of biosolid P left in soils can

be measured in various ways including by means of

plant yield, providing that all other nutrients for plant

growth are sufficiently supplied (Barrow and Camp-

bell, 1972). In this case, biosolid P remaining in soils

after N depletion may further enhance plant yields if

N fertilizer is again supplied to plants (Pascual et al.,

1999).

Nutrient availability for plant uptake depends on

soil chemical, biological, and physical conditions. Or-

ganic matter, organisms, and nutrients that remain in

biosolid-amended soils after N depletion may improve

soil conditions (Correa and Bento, 2010) and further

increase plant absorption of nutrients if N is applied on

the top of residual biosolids (Correa et al., 2005). Posi-

tive effects of residual biosolid P and organic matter in

a Spodosol have been indirectly measured through di-

fferences in plant production between chemically ferti-

lized soils (control) and soils containing residual ter-

tiary domestic sewage sludge that were applied with

urea N (Fig. 1) (Correa, 2002). Plant yields were 2–3

times higher in the Spodosol containing residual bio-

solids than in the control (Fig. 1). The application of

chemical P to the Spodosol containing residual bioso-

lids did not enhance plant yields, since P was not in

shortage in this soil after N exhaustion (Barrow and

Bolland, 1990; Correa, 2004).

Differently from the Spodosol, an Oxisol contai-

ning the same residual tertiary domestic sewage sludge

showed detrimental effect on plant yields after recei-

ving urea N (Fig. 2) (Correa, 2002). Such an effect sug-

gests shortage of available P because the Oxisol chemi-

cally fertilized at 20 mg P kg−1 (control) responded

well to increasing rates of urea N. Oxisols have a high

P-fixing capacity due to their high contents of Fe and

Al oxides (Smyth and Sanchez, 1980), which decrease

P availability to plants by means of phosphate adsorp-

tion onto soil particles (Barrow and Campbell, 1972).

As previously mentioned, tertiary sewage sludge is con-

Fig. 1 Plant yields (dry weight) at different rates of urea N applied to a Spodosol chemically fertilized at 20 mg P kg−1 (control)

and to the same Spodosol containing residual tertiary sewage sludge previously applied at rates of 1, 3, and 5 t ha−1 (Correa, 2002).

6 S. I. TORRI et al.

Fig. 2 Plant yields (dry weight) at different rates of urea N applied to an Oxisol chemically fertilized at 20 mg P kg−1 (control) and

to the same Oxisol containing residual tertiary sewage sludge previously applied at rates of 1, 3, and 5 t ha−1 (Correa, 2002).

ditioned with Al and Fe salts that can further increase

phosphate-retention capacity of soils (Whitehead et al.,

2001). Scharer et al. (2001) reported that the applica-

tion of Al and Fe oxides to a soil at 10 mg kg−1 in-

creased its P-sorption capacity by 1.6 times, with pro-

portional decrease in P availability in soil solution. As

a result, up to 90% of P in biosolids conditioned with

Al and Fe salts can not be taken up by plants (Correa,

2004; Sarkar and O’Connor, 2004). The incorporation

of tertiary sewage sludge into soils naturally rich in Fe

and Al like the Oxisol may turn the edaphic environ-

ment into a P sink. When it happens, P sources applied

at rates high enough to exceed P sorption capacity of

biosolid-amended soil can overcome shortage of avai-

lable P for plant growth (Maguire et al., 2000). A-

gricultural lime (CaCO3 or CaMg(CO3)2) amendment

can decrease soil P-sorption capacity for a while, but

P application is more effective in increasing P availabi-

lity in soils than lime amendment (Smyth and Sanchez,

1980).

Contrary to the Spodosol, plant yields respon-

ded well to the application of P fertilizer to the Oxi-

sol containing residual tertiary sewage sludge (Correa,

2002), which confirms the hypothesis of shortage of

available P in the last soil. Studies have generally re-

ported decreases in plant P uptake from soils amen-

ded with biosolids conditioned with Al and Fe salts

(Saarela, 1998; Esteller et al., 2009). Investigation on

P sorption revealed that both P isotherm slopes and

equilibrium P concentrations have unfavorably been al-

tered in the Oxisol containing residual tertiary sewage

sludge in comparison to the control Oxisol, which was

not amended with sewage sludge (Correa, 2002). Par-

ticular behaviors of biosolid P in different soils make it

difficult to generalize an optimal biosolid use for a self-

sustained positive residual effect. A specific aspect of

P dynamics in biosolid-amended soils is that the flexi-

bility of soil microorganisms influence P uptake and

immobilization rates, in particular in relation to C and

N availability by altering the C:N:P ratio (Frossard

et al., 1996b; Maguire et al., 2000). There are several

mechanisms of nutrient interactions in soils, and plant

responses to N are often dependent upon the availabi-

lity of P (Walworth and Summer, 1988).

The magnitude of P adsorption to soil particles is

also related to timing after a soil receiving phosphate

application (Burkitt et al., 2001; Whitehead et al.,

2001). Thus, biosolid P left in some soil types after N

exhaustion may not be useful for plant production after

a certain time due to soil P fixation. In other soils like

the Spodosol, plant yields were significantly enhanced

when chemical N was applied on the top of residual

biosolid P (Correa, 2002), which continues to be valu-

able for plant production. Among various minerals and

substances present in soils, Fe and Al oxides lead to the

highest P-fixing capacity due to the increasing number

of sorption sites (Celi et al., 2001; Scharer et al., 2001;

Liu Z et al., 2008). In this regard, a promising alter-

native to the use of Al and Fe salts in WWT process

is recovering N and P as struvite (MgNH4PO4·6H2O),

a slow-release fertilizer that precipitates when Mg and

lime (CaO or Ca(OH)2) are added to wastewater or

sewage sludge (Liu Z et al., 2008). This can increase

the efficiency of biosolid P use by plants (Correa, 2004)

and enhance the recycling of P from wastes in soils.

MICROBIOLOGICAL CONTROL OF P DYNAMICS

IN BIOSOLID-AMENDED SOILS

The existence of a microbial P turnover in soils is

long known and its relation with C and N cycles has

BIOSOLID APPLICATION AND GLOBAL P RECYCLE 7

been proven in the past (Johnson and Broadbent,

1952). There is a general consensus that biosolid-borne

P undergoes a slower turnover rate than P from chemi-

cal fertilizers. This is mainly due to the complexity of

biosolid matrix that becomes more recalcitrant during

decomposition. Moreover, biosolid particle allocation

within the soil aggregates also contributes to different

partition among the different biotic and abiotic pools

(Fig. 3), as hypothesized by Hens and Merckx (2001).

Soil microorganisms as well as plant roots actively

or passively release extracellular enzymes to mineralize

C, N, P, and S from complex substrates to make them

bioavailable (Nannipieri et al., 2012). Biochemical hy-

drolysis of organic phosphate esters in soils is main-

ly catalyzed by phosphomonoesterase and phospho-

diesterase, which release orthophosphate anions, the

preferentially assimilated P form by plants and soil

microorganisms. It has been reported that phosphomo-

noesterase activity in the rhizosphere is the main me-

chanism for P acquisition by plants (Gilbert et al.,

1999), catalyzing P released by a wide range of or-

thophosphate esters and anhydrides (Gellatly et al.,

1994). More complex P solubilization mechanisms, me-

diated by the release of specific secondary metabo-

lites such as polyphenols in legume plants, have al-

so been reported (Tomasi et al., 2008). In forest

soils phosphatase activity responds mainly to seaso-

nal changes of soil temperature and moisture, whereas

in arable soils, phosphatase activity mainly responds

to agricultural practices (Dick and Tabatabai, 1992)

and to the release of root exudates by crops (Renel-

la et al., 2007b). Phosphatase activity can be inhibi-

ted in soils fertilized with N, P, and K, whereas in

biosolid-amended soils, P is added together with ot-

her nutrients (e.g., C, N, and S), used as energy

sources by soil microorganisms for the synthesis of

several hydrolytic enzymes according to the economic

theory (Allison and Vitousek, 2005; Renella et al.,

2007c). This makes the major difference in P dynamics

between biosolid amendment and inorganic P fertiliza-

tion. In any case, high phosphatase activity genera-

lly observed in biosolid-amended soils does not neces-

sarily imply high P availability, as variable values of

phosphatase:microbial biomass ratio have been found

in different agro-ecosystems (Carpenter-Boggs et al.,

2003).

Increasing acid phosphomonoesterase activity has

been previously reported in agricultural soils amended

with biosolids (Dodor and Tabatabai 2003), and it is

likely related to the increases of soil microbial biomass

and activity in response to higher nutrient contents.

Different production and persistence rates of various

soil enzyme activities in different soils have also been

previously reported (Renella et al., 2007c). Biosolid-

amended soils may also undergo changes in pH-bu-

ffering capacity, which may change phosphatase acti-

vity (Renella et al., 2006). Phosphomonoesterase and

phosphodiesterase activities in biosolid-amended soils

Fig. 3 Phosphorus cycle in agricultural soils amended with chemical P fertilizers or biosolids. The thickness of the arrows and of the

boxes represent the relative importance of the pools and processes in the P cycle, respectively.

8 S. I. TORRI et al.

can be considered as indicators of potential P release

from sewage sludge because these biosolids genera-

lly contain various P forms, with a predominant pro-

portion of phospholipids (Stott and Tabatabai, 1985).

Competition between plants and microorganisms for

P in the rhizosphere mainly depends on P demand by

crops, which in turn depends on plant development

stage. Previous studies have shown that the higher the

P demand by crops, the higher the acid phosphomo-

noesterase activity, either of plant or microbial origin

(Moorhead and Sinsabaugh, 2000). Bioavailability is

the potential for a substance or molecule to be trans-

ported across the cell layer. In complex natural bodies

like soil, this pool can be determined by the use of

whole cell biosensors. Whole cell biosensors are soil-

borne bacterial strains inserted with genes producing

a detectable signal (e.g., lux for bioluminescence and

gfp for autofluorescent proteins) upon assimilation of

specific molecules (van der Meer et al., 2004). Micro-

bial biosensors responding to C, N, and P in soil have

been constructed and used to monitor the bioavaila-

ble C, N, and P pools in plant rhizosphere and bulk

soil (Kragelund et al., 1997; Darwent et al., 2003). The

use of whole cell biosensor, although not routine and

standardized methods, hold the potential to finely as-

sess P dynamics in biosolid-amended soils, where more

chemical P forms are present compared to soils ferti-

lized with inorganic P. Moreover, the development of

whole cell biosensors with multiple gene insertions and

of signaling availability for different nutrients (Koch

et al., 2001) or the use of simultaneous biosensors re-

sponding to C, N, and P (Standing et al., 2003) allows

the study of P bioavailability in function of C and N

bioavailability, which may reveal how P dynamics is

interactively influenced by C and N bioavailability. In

particular, their use may be useful to better under-

stand how C and N bioavailability influences biosolid

P mineralization/immobilization dynamics, taking in-

to account that different biosolids may widely vary in

their C:N:P ratios both at the time of application or

during decomposition (Cleveland and Liptzin, 2007).

ENVIRONMENTAL IMPLICATIONS OF BIOSO-

LID USE

In most legislation, annual application rates of bio-

solids are determined by crop N requirements. The rea-

son for this is to prevent N leaching to groundwater

(Correa et al., 2012; Al-Dhumri et al., 2013). Howe-

ver, the relatively low N/P ratio of biosolids has led

to a significant over application of P at the N-based

rate. As the amounts of P applied often exceed crop

removal (Shober and Sims, 2003; Schroder et al., 2008;

Cogger et al., 2013b), more than 95% of biosolid-borne

P remains in soils (Correa, 2004). Surplus soil P from

biosolids is not detrimental to plants. Many soils in

developed nations nowadays contain adequate to ex-

cessive P due to years of application of P fertilizers or

organic materials containing P (O’Connor and China-

ult, 2006), but soil P sorption capacity may become

saturated with time (Smith et al., 2006; Withers et al.,

2009). Maguire et al. (2000) reported increased soil P

and increased P saturation in soils receiving long-term

biosolid amendment relative to unamended soils. Past

research has shown that soils that are more saturated

with P have less capacity to retain added P, which may

increase the more labile forms of soil P, with the risk

of P loss in runoff or by leaching (Hooda et al., 2000;

Pautler and Sims, 2000). The problem arises when ru-

noff waters or subsurface flows contain environmen-

tally unacceptable contents of dissolved P forms, or

when highly P-enriched soil particles are eroded in-

to water bodies (Maguire et al., 2005). Diffuse P pol-

lution is directly associated with the development of

water body eutrophication in agricultural ecosystems

(Withers and Jarvie, 2008; Quinton et al., 2010). Al-

though both P and N contribute to eutrophication, P

is the primary agent in freshwater eutrophication be-

cause many algae are able to obtain N from the at-

mosphere (Schindler, 1977). Soluble P as low as 0.02

mg L−1 is sufficient to induce water body eutrophica-

tion (Sharpley and Rekolainen, 1997). Eutrophication

brings a series of adverse ecological and water quality

problems such as fish death, shifts in species compo-

sition, blooms of harmful algae, and hypoxia in water

body, together with the presence of toxins, taste and

odour in drinking water (Hilton et al., 2006; Lowe et

al., 2008).

Research has shown that not all biosolids have the

same potential to affect the environment when land

applied. The solubility of P in biosolids exerts a ma-

jor influence on the potential for off-site P migration

at land application sites. As mentioned above, WWT

processes govern soil P solubility because of several

factors, including biosolid treatment (especially heat

drying) and biosolid chemical composition (especially

contents of Fe, Al, and Ca). Several studies have re-

ported a relationship between low P solubility in bioso-

lids and a high content of total or amorphous Fe and Al

(Maguire et al., 2001; O’Connor et al., 2004; Krogstad

et al., 2005). Biosolid treatments that produce rela-

tively dry biosolids, like heat drying, tend to reduce

water-extractable P (WEP) (Brandt et al., 2004). Si-

nce only a small fraction of P from most conventionally

BIOSOLID APPLICATION AND GLOBAL P RECYCLE 9

produced biosolids is soluble, biosolid P should be less

likely to negatively affect the environment compared

with soluble P sources like mineral fertilizers or ma-

nures.

Even though the solubility of P in the soil increa-

ses with biosolid application rates, off-site P migration

may not necessarily increase, since a number of bin-

ding compounds incorporated through biosolids coun-

teract the leaching process. For example, Withers et

al. (2001) measured runoff P from field plots that had

previously received P from different sources, and con-

cluded that there was a lower risk of P runoff follo-

wing application of biosolid compared with other agri-

cultural P amendments at similar P application rates.

Al- and Fe-rich biosolids have been found to increase

the amorphous soil fraction, which is considered to be

a measure of the P sorption capacity of acidic soils

(Pote et al., 1996; Maguire et al., 2000). In calcareous

soils, P solubility is also influenced by Ca precipita-

tion (Pierzynski et al., 2005). Therefore, biosolids can

increase not only total P content of the soil but also

its P sorption capacity. Addition of wastes rich in Fe

and Al was also found to dramatically reduce biosolid

P leaching and runoff from high-P soils (Haustein et

al., 2000; Elliott et al., 2002b).

Different studies showed that WEP is highly cor-

related to runoff P and leachate P in manures and

manure-amended soils, and has been proposed as a

useful indicator of environmental P loss from waste-

amended soils (Kleinman et al., 2002; Brandt and

Elliott, 2003; Brandt et al., 2004). As total P varies

among organic amendments, percent WEP (PWEP

= WEP × 100/total P) is used to compare the en-

vironmentally relevant P in relation to total P. For

most biosolids, PWEP is found to be less than 5%

(Brandt et al., 2004), while Fe or Al-produced bioso-

lids have PWEP values of less than 0.5% (Brandt et al.,

2004). Conversely, BPR biosolids typically have greater

soluble P and PWEP (≥ 14%) than conventionally

produced biosolids (Brandt et al., 2004). O’Connor

and Chinault (2008) concluded that biosolid PWEP

is a very good indicator of the way that biosolid P

may affect the environment when land applied, and

proposed that biosolids with PWEP values higher

than 14% should be assumed to have a larger poten-

tial negative environmental impact than biosolids with

PWEP values less than 14%.

Until recently, P has been thought to be so strong-

ly bound to the soil matrix that its vertical movement

through the soil profile is insignificant (Kostyanovsky

et al., 2011; Oladeji et al., 2013). Since most soils have

an appreciable P-sorbing capacity, P that may move

down the soil profile generally becomes fixed in the

subsoil. Hence, P leaching is not considered an im-

portant P loss mechanism (Miller, 2008). Therefore,

numerous studies concluded that P vertical movement

through the soil profile in biosolid-amended soils was

negligible, despite the high rates of P applied or soil

texture. However, over application of P to soils with

low P sorption capacity may significantly increase P

vertical movement and leaching. Leaching of P from

organic amendments may occur in both organic and

inorganic forms (Eghball et al., 1996). Complexation

of P with mobile organic compounds may favour the

deep transport of organic forms of P, even through lay-

ers with a great P adsorption capacity. In a column

experiment using a fine sandy soil amended with six

conventional treated biosolids at N-based rates, P lea-

ching was less than 1% of P applied, and not statistical-

ly different from unamended soils. In contrast, 21% of

the P applied was found to leach in columns amended

with TSP (Elliott et al., 2002a). Rydin and Otabbong

(1997) leached 35 mm of water through soils amended

with either Fe- or Al-treated biosolids and found that

less P was released from Fe-treated biosolids compared

with Al-treated biosolids. When only biological treat-

ment processes are involved, biosolids are usually re-

ported to have a relatively high risk of P leaching than

soils amended with biosolids stabilized with high levels

of Fe or Al (Kyle and McClintock, 1995). This varia-

tion is most likely due to differences in solubility in the

forms of inorganic P resulting from different WWTPs.

Thermal treatment of biosolids was also found to sig-

nificantly reduce P leaching in sandy soils (O’Connor

et al., 2002), because heating increases the rate of reac-

tion of simple, readily dissolvable phosphate minerals

to more complex, less soluble forms.

Runoff losses of P may occur in particulate and

soluble P forms. Particulate P is associated with soil

particles, such as minerals or organic matter. Runoff

of particulate P may be decreased through different

management practices (Kleinman et al., 2011; Dodd

and Sharpley, 2015), but soluble inorganic P loss is of

concern, especially in low P-retaining soils (McDowell

et al., 2004; Shober and Sims, 2007). Penn and Sims

(2002) noted that runoff P is very high (0.064 mg L−1)

from soils amended with BPR biosolids, followed by Fe

and lime-treated biosolids (0.039 mg L−1), no-Fe and

no-lime biosolids (0.014 9 mg L−1), and Fe-treated and

no-lime biosolids (0.002 mg L−1) at equal rates of to-

tal P (200 kg ha−1). The reason for this situation is

that P amendments which do not add P-binding ele-

ments (e.g., BPR) can be expected to increase P sa-

turation, reduce P-binding strength, and release more

10 S. I. TORRI et al.

P to runoff (Holford et al., 1997; Siddique and Robin-

son, 2003). When only biological treatment processes

are involved, biosolids are usually reported to have a

relatively high risk of off-site P migration than those

stabilized with high levels of Fe or Al (Kyle and Mc-

Clintock, 1995). Field studies of White et al. (2010)

have shown that runoff P for the soils amended with

Fe-treated biosolids is not significantly different from

that for the unamended control soil despite biosolid

application rates. The soils amended with lime-treated

biosolids produce the largest runoff P, the soils amen-

ded with Fe and lime-treated biosolid are intermediate,

and those amended with Fe-treated biosolids are the

lowest. These have been attributed to the dissolution

of calcium-bound P (Ca-P) species in acidic soils af-

ter land application of biosolids (Leytem et al., 2004;

White et al., 2010). Most research has shown that the

addition of metal salts at the WWTP reduces solu-

ble P losses by runoff (Penn and Sims, 2002; Agyin-

Birikorang et al., 2008; Alleoni et al., 2008). Elliott et

al. (2005) reported that with additions of Fe and/or Al

during WWT processes, like heat drying, runoff P los-

ses produced are not statistically different between the

amended and unamended soils. Other researchers re-

ported that some biosolid-amended soils produced less

runoff P losses than the unamended soils (Brandt and

Elliott, 2003; O’Connor and Elliott, 2006).

A peculiar environmental behaviour of soil P dy-

namics is the so-called P leaching breakpoint, first

observed in the long-term Broadbalk Experiment at

Rothamsted, UK (Heckrath et al., 1995). The P lea-

ching breakpoint indicates an abrupt change occurring

in the Olsen P fraction when it is in the range of 21–104

mg P kg−1. The occurrence of a P leaching breakpoint

has been confirmed for other soils under various mana-

gement practices (Brookes and Hesketh, 1998; Jordan

et al., 2000). To our knowledge, the existence of a P

leaching breakpoint in biosolid-amended soils has not

been studied. This aspect may be important from the

perspective of utilizing biosolid P, as the higher or-

ganic matter content of biosolid-amended soils should

theoretically saturate P sorption sites, leading to po-

tentially greater P losses compared to those of the soils

amended with chemical fertilizers.

CONCLUSIONS AND PERSPECTIVES

Phosphate rock is a finite, non-renewable resource,

and its reserves are progressively becoming scarce. Re-

cycling P from biosolids is a valuable feedstock for

agronomic purposes to enhance and sustain society,

and represents the best environmental option so far.

However, land application of biosolids is becoming in-

creasingly constrained by the amounts of P addition in

sensitive agronomic scenarios. It is generally accepted

that leaching of P from biosolid-amended soils is mini-

mal. However, the risk of soluble inorganic P transport

in surface runoff after land application of biosolids is

of major concern. The WWT processes clearly influ-

ence differences in soil P solubility and soil P speci-

ation after land application of biosolids. In sensitive

scenarios, Fe- or Al-treated biosolids reduce the risk

of P transport. However, if runoff P is not a major

concern and biosolids are primarily applied to provide

available P to crops, the standard BPR process or a

process that involves the addition of lime instead of

Fe and Al oxides may be adequate. In all cases, it is

critical to control sources of nonpoint P pollution of

surface- and groundwater. While in natural soils, the

phosphatase activity likely plays an important role in

P mineralization and phytoavailability, other microbio-

logical and biochemical activities likely play predomi-

nant roles in P mineralization and fate. The use of

whole cell biosensors specifically signalling to P up-

take by soil microorganisms is a promising biotech-

nology for the assessment of the P bioavailability in

soil which can improve understanding of P released by

biosolid application. Further research on P forms in

the various biosolids, the use of biotechnologies for the

assessment of the P bioavailable fractions such as the

whole cell biosensors, and the analysis of genetic plant

responses to soil biosolid amendment can improve the

understanding of potential P uptake by crops and opti-

mal use of P-rich biosolids for sustainable agriculture.

Harmonization of the regulation on the use of bioso-

lids in agriculture, currently mainly based on N and

pollutant contents, may also contribute to a better P

balance in agriculture.

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