Biosolids application to agricultural land: a contribution to global phosphorus recycle,

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 to
Global Phosphorus Recycle: A Review
Silvana Irene TORRI1,∗
, Rodrigo Studart CORRÊA2
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 Bras´ılia-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, Corrêa 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-
c´ıa-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; Mart´ınez et
al., 2007; Torri, 2009; Ahumada et al., 2014; Jordán 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

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 (Sánchez-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-

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 (Corrêa, 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-
rêa 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 (Corrêa and Bento, 2010) and further
increase plant absorption of nutrients if N is applied on
the top of residual biosolids (Corrêa 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) (Corrêa, 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; Corrêa, 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) (Corrêa, 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 (Corrêa, 2002).

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 (Corrêa, 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 (Corrêa,
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 (Corrêa,
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 (Corrêa, 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 (Corrêa, 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 (Corrêa, 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

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.

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
(Corrêa 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 (Corrêa, 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

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

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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