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IV. Impact of Phosphorus on the Aquatic Environment

IV. Impact of Phosphorus on the Aquatic Environment

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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).

A.



SOLUBLE

PHOSPHORUS



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.

B. PARTICULATE

PHOSPHORUS

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



313



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).



c.



BIOAVAILABILITY

OF PHOSPHORUS



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



3 14



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



Table N

Percentage Bioavailability of Sediment P Transported in Several Lake Tributaries Draining

Agricultural Watersheds and in Deposited Lake Sediments



Bioavailabilitf

Reference



Location



Suspended sediment in tributaries

Dorich et al. (1980)

Indiana



Description

Agricultural



De Pinto el al. (1981)



Great Lakes



Agricultural



Logan et al. (1979)



Lake Erie



Agricultural



Central Canada

Lake Ontario



Prairie Lakes

Postglacial

Glacial

Basin

C. L. Memphremagog

S. L. Memphremagog

Riviere-du-sud

Rijuland water

Calcareous

Noncalcareous

Bottom sediments



Deposited sediment

Allan and Williams (1978)

Bannerman et al. (1975)

Carigan and Kalff (1980)



Quebec



Klapwijk et al. (1982)

Sagher et al. (1975)



Netherlands

Wisconsin



Williams et 01. (1980)



Great Lakes



Procedure



(Qo)



Bioassay

NH, F

NaOH

HCl

Bioassay

NaOH

CDB*

NaOH

CDB



21

9

8

4

0-4

4-38

9-27

14-42

29-56



CDB

NaOH

NaOH

NaOH

Resin

Resin

Resin

Bioassay

NaOH

NaOH

NTA~

NaOH

Resin



14-37

30-60

2-8

13-18

8

25

19

0-41

60-95

80-85

30

27

21



Total P

Wkg)



0.2-0.7



0.5-1.2

0.5-1.2



0.5-1.3

1 .O-1.5

0.9-1.0

0.8-0.9

0.8-1.2

0.4-4.8

0.6-3.9

0.4-1.4



'Tercentage total particulate P bioavailable.

bCDBand NTA representcitrate-dithionate-bicarbonateand 0.01 Mneutralized nitdoacetic acid extractableP, respectively.



316



A.



N. SHARPLEY AND R.



G. MENZEL



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).



D. ARTIFICIAL

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



317



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

mesotrophic.



V.



CONCLUSIONS



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



318



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



319



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.



ACKNOWLEDGMENT

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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