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IV. Impact of Phosphorus on the Aquatic Environment
A. N. SHARPLEY AND R. G.MENZEL
biochemical, and physical processes. Although soluble P is immediately
available for algal uptake, particulate P may provide a long-term source of
available P to algae growth through desorption to the surrounding lake
water (Bjork, 1972; Larsen et al., 1975; Cooke et al., 1977). Thus, the processes controlling the bioavailability of particulate P must be considered in
designing programs to control accelerated eutrophication. Soluble and particulate P may be removed from the biotic zone by the natural processes of
phytoplankton uptake and deposition. The process of accelerated
eutrophication has been temporarily reversed in several eutrophic and
hypereutrophic lakes by the inactivation of biologically available P with the
addition of alum (Peterson et al., 1973; Cooke et al., 1978).
The most available form of P to algae in the aquatic environment is soluble P (Vollenweider, 1968; Bartsch, 1969). Walton and Lee (1972) reported
that soluble P was essentially 100% available, using algal assay procedures
and a variety of waters. A number of investigators, however, have found
that soluble P as measured by the molybdate method (Murphy and Riley,
1962) is not completely available to support algal growth (Rigler, 1968;
Lean, 1973a,b: Dick and Tabatabai, 1977; Stainton, 1980). This results
from a possible reduction in condensed phosphates, hydrolysis of organic P
compounds, and reaction with arsenate during analysis, all of which will
contribute to an overestimation of the true soluble P concentration. This
discrepancy is relatively great for waters of low P concentration, such as are
normally found in lakes, while the percentage error is much lower with concentrations found in streams, rivers, or wastewater discharges. Lee et al.
(1979) suggested that from a lake management point of view, the discrepancy at low soluble P concentrations is of no major consequence as P control
programs must be directed at sources of high concentrations.
Boyd and Musig (1981) observed that planktonic communities in samples
of water from fish ponds absorbed an average of 41070 of 0.30 mg/liter additions of soluble P within 24 hr. Over a longer period of time (2 weeks), these
concentrations declined to 10% of that originally present due to the added
removal of P by sediment.
In oligotrophic and sometimes in eutrophic waters where soluble P concentrations are depleted by vigorous algal growth, concentrations may be as
low as 0.001 mg/liter (McColl, 1972). Under these conditions, P may be
desorbed from the suspended or deposited sediment material. In fact, Bannerman et al. (1975) calculated that approximately 10% of the external P
THE IMPACT OF SOIL AND FERTILIZER PHOSPHORUS
loading of Lake Erie (1.3 x lo7 kg/ha) resulted from the desorption of P
from lake sediments. Furthermore, several studies have reported that particulate P can support biological growth, even though the soluble P concentration is low (Healy and McColl, 1974; Golterman, 1977; Allan and
Williams, 1978). Consequently, the capacity of sediment entering lakes to
supply or remove P is important in an evaluation of the importance of P in
the aquatic environment.
The processes involved in the sorption and desorption of P by sediment
material in a lake are analogous to those occurring in stream flow discussed
earlier. The direction of P movement will be governed by the soluble P concentration of the lake water and the desorbable P content or EPC, of the sediment
material. The rate and extent of P interchangebetween sediment P and the surrounding lake water is controlled by the forms of P contained in the sediment
and soluble P concentration of the interstitial water. The forms of P contained
in the sediment have been reviewed extensively by Syers et al. (1973b). The
potentially mobile forms are P sorbed on hydrous Fe and Al oxides and CaCO,.
In addition to the chemical mobility of sediment material, its physicalmobility will also affect the interchangebetween particulate and soluble P, and subsequently its bioavailability. The physicalmobility of sedimententeringa lake will
be a function of its texture and lake water temperature and turbulence. The
relative densities and temperatures of the inflow and lake water will determine
whether sediment enters the surface or bottom waters of the lake. As it enters
the lake its turbidity may reduce the depth of the photic zone. Coarse-textured
sedimentswill settlerapidly and be available to algae in the photic zone for short
periods only. In contrast, fine-texturedsedimentswill remain in the photic zone
for a longer period of time. The bioavailability of this sediment will be further
increased by the fact that it will be enriched in P compared to coarser material.
Removal of fine-textured sediments from the photic zone may be enhanced by
bioflocculation in the presenceof certain algae (Avnimelechand Menzel, 1984).
The mobilty of particulate P can increase if the sediment settles from an
oxic photic zone into a deoxygenated hypolimnion. The desorbed P can
then be redistributed during periods of lake turnover. In a study of the P
dynamics of two shallow hypereutrophic lakes in Indiana, Theis and McCabe (1978) found that the soluble P concentration of lake water was reduced by sorption during oxic periods and increased by desorption during
anaerobic periods. The increased mobility of particulate P under anaerobic
compared to aerobic conditions is attributed to a reduction of Fe(II1) to
Fe(I1) (Li et al., 1972; Syers et al., 1973b; Patrick and Khalid, 1974).
Due to the importance of sediment as the major source of P entering the
aquatic environment from agricultural land and its ability to sustain algal
A. N. SHARPLEY AND R. G . MENZEL
growth, several methods to estimate the bioavailability of this P source have
been proposed. The availability of P to algae can be determined by an algal
culture test (EPA, 1971). However, more rapid chemical extraction procedures, which simulate removal of P by algae, have been proposed for the
routine determination of particulate P bioavailability (Dorich et al., 1985).
Chemical extractants that have been used to measure the bioavailability of
particulate P are NaOH (Sagher et al., 1975, Golterman, 1976; Cowan and
Lee, 1976; Armstrong et al., 1979; Logan et al., 1979); NH,F (Porcella et
al., 1970; Dorich et al., 1980); anion exchange resins (Wildung and
Schmidt, 1973; Cowan and Lee, 1976; Armstrong et al., 1979; Huettl et al.,
1979); and citrate-dithionite-bicarbonate (CDB) (Logan et al., 1979). It is
suggested that the weaker extractants and short-term resin extractions
represent P that could be utilized by algae in the photic zone of lakes under
aerobic conditions. In contrast, the more severe extractants (CDB) represent P that might become available under reducing conditions found in the
anoxic hypolimnion of stratified lakes.
Caution must be exercised, however, in relating P bioavailability of sediment material determined by these chemical extractions and the potential of
the sediment to increase algal growth (Lee et al., 1979; Sonzogni et al.,
1982). In turbid stratified lakes the surface photic zone may be relatively
thin compared to the mixed layers above or below the thermocline. In addition, suspended sediment material often contains large amounts of silt-sized
aggregates of clay, which will settle more rapidly from the photic zone than
smaller particles, possibly reducing the actual availability of particulate P.
Consequently, bioassays may produce erroneous estimates of available particulate P unless the physiochemical properties of the waterbody and sediment are considered in determining the appropriate bioassay to be used.
Although these above chemical extraction procedures have identified
which particulate P fractions can be utilized by algae, there is no evidence
that all of the chemically extracted P is algal-available. Thus, Hegemann et
al. (1983) suggested that a quantitative assessment of algal-available particulate P will depend upon the development of long-term (> 100 day) algal
assay procedures. The bioavailability of P attached to suspended sediment
transported in tributaries to lakes and of sediment deposited in lakes is summarized in Table IV for several studies. It is evident that a large variability
in the bioavailability of sediment P exists, which reflects the dynamic nature
of the physiochemical processes governing the transport, P mobility, and
deposition of eroded soil material. Phosphorus associated with suspended
sediment can be considered to be of short-term bioavailability due to
sedimentation from the biotic zone.
In contrast, P associated with deposited sediments is potentially
bioavailable for a much longer period of time. Wildung et al. (1974)
reported that the P content of the sediment in several lakes in Oregon was
Percentage Bioavailability of Sediment P Transported in Several Lake Tributaries Draining
Agricultural Watersheds and in Deposited Lake Sediments
Suspended sediment in tributaries
Dorich et al. (1980)
De Pinto el al. (1981)
Logan et al. (1979)
C. L. Memphremagog
S. L. Memphremagog
Allan and Williams (1978)
Bannerman et al. (1975)
Carigan and Kalff (1980)
Klapwijk et al. (1982)
Sagher et al. (1975)
Williams et 01. (1980)
'Tercentage total particulate P bioavailable.
bCDBand NTA representcitrate-dithionate-bicarbonateand 0.01 Mneutralized nitdoacetic acid extractableP, respectively.
N. SHARPLEY AND R.
directly related to the biological productivity of surface waters and served as
a significant source of P to these waters, supporting increased biological
growth. Carignan and Kalff (1980) found that submerged macrophytes
depended overwhelmingly on sediments for their P supply. Even under
hypereutrophic lake conditions, sediments contributed the major proportion (72010) of P utilized during growth. It has been suggested, moreover,
that these aquatic plants may supply P to overlying waters by excretion during growth and upon senescence (Carignan and Kalff, 1980).
The water renewal time of a lake plays an important role in the dynamics
and extent of P exchanges in a lake. With a short residence time, outflow of
water from a lake can be a more important route for P removal than
sedimentation. When the residence time of a lake exceeds a few months,
most of the P inflow is retained in the lake sediments. Because of this process, impoundments and small lakes have been used as efficient traps
(especially for particulate P) to improve downstream water quality (Rausch
and Schreiber, 1977). It is apparent, however, that the amounts of P stored
in lakes can build up to unacceptable levels, resulting in a permanent
deterioration in water quality. In fact, a reduction in the external load of P
upon the highly eutrophic Lake Trummen in Sweden did not bring about
the desired improvement in water quality until the upper layers of the P-rich
sediment were removed (Bjork, 1972).
REMOVAL OF PHOSPHORUS FROM LAKES
In order to control or reduce the increased biological productivity of
lakes and impoundments, the inputs of P must first be reduced. The diversion of P inputs, however, does not always bring about a prompt and sufficient reduction in lake water concentration, due to internal recycling from
P-rich sediments (Larsen et al., 1975; Cooke et al., 1977). A reduction in the
P concentration of lake water and the inactivation of the recycling
mechanisms may be brought about by chemical amendments (Jernelov,
1970; Peterson et al., 1973; Cooke et al., 1978; Kennedy, 1978). Several
points need to be considered in the restoration of lake water quality by P
precipitation and inactivation. These include the chemicals used, dosage, effect of the additives on benthic fauna, and method and time of application.
The most commonly used chemicals are aluminum sulfate and sodium
aluminate, due to the stability of flocculated Al hydroxides with redox
changes. The removal of P is brought about by precipitation of AlPO,, by
coagulation or entrapment of P-containing particulates, or by sorption of P
on the surfaces of A1 hydroxide polymers (Recht and Ghassemi, 1970;
Eisenreich et al., 1977).
The maximum dosage of A12(S0,)3 for the long-term control of P cycling
may be determined by Al,(S04)3addition to lake water samples until the
THE IMPACT OF SOIL AND FERTILIZER PHOSPHORUS
dissolved Al concentration reaches 0.050 mg AlAiter (Kennedy, 1978), a
concentration Everhart and Freeman (1973) found to be nontoxic to fish.
Very little direct laboratory or field evidence on the effect of Al on the
aquatic biota exists, although several studies have shown no apparent effect
on fish (Kennedy and Cooke, 1974; Bandow, 1974; Sanville et al., 1976) or
benthic invertebrates (Narf, 1978) following full-scale lake treatments.
A predetermined amount of A12(S04),is applied as a slurry from the lake
surface if P removal from the epilimnion is required. If control of P release
from sediments is required then application to the hypolimnion is necessary.
As d2(so4)3removes dissolved organic P inefficiently (Browman et al.,
1973; Eisenreich et al., 1977), applications should be made in early spring
when the major proportion of P in lake water is inorganic (Browman et al.,
1977; Eisenreich et af., 1977). The continued presence of organic P may be
significant, as Heath and Cooke (1975) observed that certain nuisance bluegreen algae can produce a phosphatase enzyme under P-limiting conditions,
that is capable of mineralizing organic to inorganic P at rates sufficient to
support algal blooms. Application time will not be critical for treatment of
P desorption from lake sediments. However, the relative importance of lake
sediments as a P source should be assessed prior to Al,(SO4), application.
For example, lakes receiving substantial inputs of clay in addition to P may
contain sediments with high sorption capacities for P.
The application of A12(S04),to just below the surface of Horseshoe Lake,
Wisconsin, resulted in a significant decrease in the P content of both the
epilimnion and hypolimnion (Peterson et al., 1973). Prior to application,
the lake had experienced algal blooms and fish kills which were partially attributed to agricultural inputs of P. Born (1979) observed that although
hypolimnion P increased slightly each year after application, it never reached pretreatment levels, thus giving approximately 8 years of control. The
hypolimnetic application of Al,(S04), to the eutrophic West Twin Lake,
Ohio, resulted in an 88% reduction in total P concentration of the lake
water (Kennedy, 1978). Continued water quality monitoring by Kennedy
(1978) indicated that the layer of Al(OH), deposited on the sediments reduced P release to overlying waters by 98%. Three years later the lake was
Fertilizer P use presents no direct problem to the terrestrial environment.
The use of P fertilizer is essential to maintain adequate crop production for
an ever-increasing population. Its application can reduce the nutrient
enrichment of surface waters by establishing an increased vegetative cover
A. N. SHARPLEY AND R. G . MENZEL
on eroding soils. These benefits to the environment must be considered
along with a potentially detrimental indirect effect of fertilizer P use. Heavy
metal and radionuclide contaminants not removed during fertilizer
manufacture may xcumulate in the soil. No threat to human health has
been reported at the present time, although the continued application of impure fertilizer materials to acid soils may lead to problems with crops
susceptible to contaminant uptake, especially that of cadmium.
Most of the public and scientific attention regarding environmental effects of P fertilizer has been focused on the aquatic environment due to the
role of P in increasing the biological productivity of lakes and impoundments. Considerable research has been conducted to quantify the losses of
soil and fertilizer P from various land management practices. However, we
are still unable to relate P inputs to a lake or impoundment to a quantitative
description of water quality. Furthermore, the effect of P concentration on
algal growth receives continued attention, while little information is
available on how lake macrophytes are affected, even though macrophytes
present a more serious economic problem than algae in many lakes.
Research should be directed towards improving the partitioning models
for soluble and particulate P transport in runoff and in lakes and impoundments. This should focus on the mechanisms of exchange between desorbable or labile P and solution and methods to routinely quantify the
amounts of desorbable or bioavailable P on various materials. With the accumulation of P at the soil surface under conservation tillage practices, existing soil P test procedures may need to be reevaluated. This may include
changes in soil sampling frequency, timing, and depth, in addition to the
use of chemical extractants capable of removing easily mineralizable
organic P from the surface plant residue built up.
As the crop canopy can contribute a major proportion of the soluble P
transported in runoff, surface soil and plants must be considered as a continuum and the pool of desorbable P in both soil and plant material determined. The measured size of this pool will depend upon the experimental
conditions of analysis, therefore extraction mediums, so1ution:soil ratios,
and contact times relevant to either the terrestrial or aquatic environments
must be used. In the case of the aquatic environment, the fact that the
desorbed P can be continuously removed from the system by algal growth
must be considered.
In the light of research on the kinetics of P exchange between desorbable
P and solution in the terrestrial and aquatic environments and during
transport from the terrestrial environment, more accurate and widely applicable models simulating P transport from watersheds can be expected.
This information should be used to improve the prediction of both the
amounts and forms of P transported into lakes and impoundments. These
models can then be used as tools to aid management decisions to reduce P
THE IMPACT OF SOIL AND FERTILIZER PHOSPHORUS
loss in runoff and at the same time to increase crop yields to maintain adequate food production for an increasing population. In addition, these
models may also identify areas of further research.
The authors wish to acknowledge the pioneering phosphorus and water quality research of
Dr. J. C. Ryden, who died suddenly May 3, 1986, in London, England.
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