NUTRIENT CYCLING IN THE GREAT LAKES::

A SUMMARIZATION OF FACTORS
REGULATING THE CYCLING OF PHOSPHORUS

Donald Scavia and Russell Moll
Blditors

University of Michigan - Great Lakes Environmental Research Laboratory

Cooperative Program in
Great Lakes Longterm Effects Research

Great Lakes Research Division Special Report No. 83

December 1980

TABLE OF CONTENTS

Page
INTRODUCTION « iv

CHAPTER

I. Measurements of phosphate-phosphorus in lake water 1

Stephen J. Tarapchak

II. Chemical factors controlling phosphorus cycling in lakes ..• 13

Wayne S. Gardner and Brian J. Eadie

III • Bacterial dynamics and phosphorus cycling • 35

Russell Moll and Mark Brahce

IV. Lower food web dynamics

Cycling of P among phytoplankton, herbivorous zooplankton,

and the lake water 49

John T. Lehman

Phosphorus cycling through zooplankton as influenced

by their diurnal migration, predatory feeding, and

population dynamics 69

James A. Bowers

V. Fish dynamics and phosphorus cycling in lakes 81

James F. Kitchell

VI. The biological component of phosphorus exchange and cycling in

lake sediments • 93

Thomas F. Nalepa, David S. White,

Catherine M. Pringle, and Michael A. Quigley

VII . The role of macrophytes in phosphorus cycling Ill

Kwei Lin

VIII. Conceptual model of phosphorus cycling 119

Donald Scavia

111

INTRODUCTION

The Laurentian Great Lakes of North America contain nearly 20% of the
world's lake and river waters and constitute one of the world's largest
freshwater resources. Large concentrations of people and industry surround
these inland seas. More than one-third of Canada's population and one-eighth
of the United States' population lives within the Great Lakes basin. The
region accounts for one-fifth of the annual U.S. GNP and one-half of the annual
Canadian GNP. Nearly all human activities in the region are associated with
the lakes. Those activities include uses for municipal, agricultural, and
industrial water supplies; recreation; mining; fishing; and transportation.
The Great Lakes also have been used as a convenient disposal site for the
wastes of the highly industrial society along the shoreline.

The careless use of the Great Lakes as a disposal site for both toxic and
nutrient wastes has diminished the resource value of the lakes, particularly
the freshwater fisheries (Smith 1968, 1972). In response to the deteriorating
water quality in the Great Lakes , many United States and Canadian agencies have
been wrestling with the problem of lake restoration (International Joint
Commission 1977). While significant strides have been made in reversing
deterioration on a lake-wide basis, it has been recognized that current
policies will not likely provide satisfactory improvement in bays and nearshore
regions (Thomas et^ al^» 1980). These areas, which are highly impacted by
pollution and readily perceived by many people, require additional attention.
Subsequent to existing management policies, which are oriented toward relieving
background conditions, new policies must address these localized areas.

Cognizant of this requirement and based on the premise that more detailed
information is needed to support these second-order policies, the Great Lakes
Environmental Research Laboratory is designing long-term research programs to
focus on these problems. Prior to design of the research plan, the state of
our understanding of the biological, chemical, and physical processes that
interact with contaminating substances must be documented.

The purpose of this report is to address the state of our understanding of
one aspect of Great Lakes ecosystem function, the cycling of phosphorus.
Previous research has identified phosphorus, in particular, as one of the key
elements in causing water quality degradation and/or eutrophication in lakes
(Dillon and Rigler 1974, Schindler 1978). The rationale for further addressing
phosphorus as a polluting substance originates from emphasis on more spatially
localized problems, as discussed above. Whereas previous analyses at lake-wide
average and annual scales could ignore detailed aspects of internal recycling,
present analyses of more localized problems on shorter time scales require
improved information on those cycling mechanisms. This report represents the
initial step in addressing the research problem on these scales. The
information presented below is a review of the state-of-the-art in phosphorus
cycling in freshwaters.

The cycling of phosphorus, although of interest ultimately to regulatory
agencies, is basically an ecological problem and has been addressed herein
through an ecosystem approach. This approach recognizes that essential
nutrients affect all components of the aquatic food web both individually and
in a composite fashion. The chapters below address each segment of the
ecosystem, beginning with chemical species and fractions of phosphorus and
moving through bacteria, phytoplankton, zooplankton (herbivores and

iv

carnivores), fish, macrophytes , and benthos. Each chapter is intended to be
compreheinsive, but not exhaustive. Each chapter is quantitative when possible,
although many areas of phosphorus cycling are insufficiently understood for
extensive quantification. The movement of phosphorus by predominately physical
means has not been addressed in this report, but is considered elsewhere
(Lick 1979).

In the final chapter, a framework for understanding the composite movement
of phosphorus through the lake ecosystem is presented. This chapter begins
with a review of current phosphorus models j, and then develops a conceptual
model of phosphorus cycling based on a synthesis of available information drawn
primarily from the other chapters. The rationale for this model is to place
the importance of each component of the food web into perspective in terms of
overall phosphorus cycling, and to emphasize those areas where basic
information is lacking.

Before the reader begins this report, the intent in preparing this
document should be recalled. As mentioned above, this report is intended to be
comprehensive, but not exhaustive; it is intended to be factual and
informative, but not speculative; and it is intended to serve as a basis for a
high quality research program exploring and identifying the major pathways and
controls of phosphorus cycling in the nearshore regions of the Great Lakes.

REFERENCES

Dillon, D. J. and F. H. Rigler. 1974. A test of a simple model predicting
the phosphorus concentration in lake water. J[. Fish . Res . Board Can .
31: 1771-1778.

International Joint Commission. 1977. New and Revised Great Lakes water
quality objectives . Vol. 1. Windsor, Ont.

Lick, W. 1979. The transport of contaminants in the Great Lakes . Report

to Great Lakes Environmental Research Laboratory, Ann Arbor, Mi., 53 pp.

Schindler, D. W. 1978. Factors regulating phytoplankton production and

standing crop in the world's f reshwaters . Limnol . Oceanogr . 23: 478-486.

Smith, S. H. 1968. Species succession and fishery exploitation in the
Great Lakes. J_. Fish. Res. Board Can . 25: 667-693.

1972. The future of salmonid communities in the Laurentian

Great Lakes. £. Fish . Res. Board Can . 29: 951-957.

Thomas, N. A., A. Robertson, and W. C. Sonzogni. 1980. Review of control
objectives: New target loads and input controls, pp. 61-90.
In: (R. C. Loehr, C. S. Martin, and W. Rast, eds.) Phosphorus management
strategies for lakes. Ann Arbor Science Publ., Ann Arbor, Mi.

VI

CHAPTER I

MEASUREMENTS OF PHOSPHATE -PHOSPHORUS
IN LAKE WATER

Stephen J. Tarapchak

Great Lakes Environmental Research Laboratory

National Oceanic and Atmospheric Administration

Ann Arbor, Michigan

INTRODUCTION

Reviews on methods for measuring phosphate-phosphorus (PO4-P) in natural
waters have been prepared by Harwood and Eattingh (1973), Kimerle and Rorie
(1973), Chamberlain and Shapiro (1973), and Rigler (1973). The purpose of this
chapter is to review information on chemical methods for measuring PO4-P, to
describe errors or inconsistencies that can be encountered in performing
analyses, and to summarize the importance of interference caused by the
hydrolysis of organic compounds or by release of PO4-P from particles during
conventional molybdentjm blue analysis. The relationship between bioassay
methods for measuring "biologically available P" and analytical PO4--P
measurements is also summarized.

The basic method for estimating PO4-P concentrations in solution was
introduced by Osmond (1887) and applied to analysis of lake-water samples by
Deniges (1920) and Atkins (1923). The basic reaction involves complexing PO4-P
with molybdate under acid conditions to form phosphomolybdic acid (a yellow
heteropoly acid), which is subsequently reduced by stannous chloride to form a
blue complex. The method was improved by Harvey (1948) by adjusting for an
optimuia ratio of acid normality to molybdate concentration in order to form a
stable color and to detect PO4-P at concentrations approaching 1.0 yg P/L.
Since that time, the method has been continuously modified. Olsen (1967), for
example, lists over 100 variants of the method that involve alterations in acid
normality, molybdate concentration, color development time, and procedures to
eliminate interferences or to reduce hydrolytic potentials.

With gradual improvements in analytical techniques and equipment and an
understanding of the phosphorus fractions in lake water, the terminology
associated with phosphorus analysis has evolved over time (see Olsen 1967).
Strickland and Parsons (1965) give the following scheme for the different
fractions of P in natural waters (from Rigler 1973):

Soluble reactive phosphorus (SRP) refers to the value obtained when
me^mbrane- filtered water is analyzed by one of the variants of the
molybdenum blue technique. This term implies neither that the
orthophosphate measured was in solution before addition of the reagents
nor that the intensity of the blue color is exclusively a function of
orthophosphate concentration rather than that of interfering ions. When
"orthophosphate phosphorus" (PO4-P) is used, it will not refer to the
results of chemical analyses but to free orthophosphate in solution, the
concentration of which is assumed to be as yet unmeasurable in the
trophogenic zone of most lakes.

Soluble phosphorus (SP) refers to the value obtained when
meimbr an e- filtered (0.45-ym pore size) water is analyzed after being
digested with an oxidizing acid solution.

Soluble unreactive phosphorus (SUP) is the difference between SP
and SRP.

Total phosphorus (TP) is obtained by analyzing whole lake water after
acid digestion. It is assumed that the values obtained by this technique
are indicative of the true phosphorus content of the sample.

Particulate phosphorus (PP) is the total phosphorus minus soluble
phosphorus.

As the chemical composition of different phosphorus fractions in lake

water becomes better understood in the future, a more precise and descriptive
terminology undoubtedly will evolve. Throughout this and subsequent chapters,
the above terminology is used.

TOTAL, PARTICULATE, AND TOTAL SOLUBLE PHOSPHORUS

TP, PP, and SP are determined routinely by digesting samples of lake water
and then measuring molybdate-reactive PO4-P by any one of several molybdenum
blue methods (A.P.H.A. 1976, Strickland and Parsons 1972, Golterman and Clymo
1971). Digestion in H2SO4, HNO3, persulfate, or perchloric acid under pressure
are common procedures for oxidizing samples. PP normally is estimated by
subtracting SP from TP. These digestion procedures, however, are not adequate
for some lake waters where highly stable P compounds or minerals containing
PO4-P are present. In such cases perchloric acid, a very strong oxidant, is
used in place of acid or persulfate.

METHODS

The basic colorimetric method for SRP analyses involves the addition of
ammonium molybdate to an acidified sample (usually 0.1 to 0.4N) to form
12-phosphomolybdic acid which is subsequently reduced (generally with stannous
chloride) to form a blue color. After 6 to 12 min is allowed for color
development, the optical density of samples and standard solutions of known
PO4-P concentrations are measured (at a wave length near 650 m) in a
spectrophotometer. SRP concentrations in samples are obtained from a plot of
optical density (corrected for color in reagent blanks) of PO4-P standards
versus known PO4-P concentrations. Until the mid 1960s, Harvey's technique in
one form or another was the time-honored method for measuring SRP.

The ascorbic acid method, developed by Murphy and Riley (1958, 1962), has
gained wide-spread popularity within the last 10 years. The method employs
ascorbic acid for rapid color development. The principal advantages of the
technique are that color development is stable for up to 24 hr or more and all
reagents are added from a single "mixed solution," whereas the phosphomolybdate
complex formed by Harvey's method is stable only for 20 to 30 min after the
reductant is added and acid molybdate and reductant are added individually to
samples. Both methods are precise (+0.10 yg P/L) and have detection limits of
ca. 1.0 lig P/L using modern spectrophotometers and cuvettes with a light path
of 5 to 10 cm in length. Since SRP concentrations in oligotrophic lakes often
are at or below the detection limits of conventional methods, extraction
techniques have been developed to increase sensitivity. For example, Stephens
(1963) extracted reduced phosphomolybdate by adding reagents of the ascorbic
acid method to filtered samples and extracted the reduced complex into
isobutanol. Color is stable for 2 hr, the limit of detection is 0.2 yg P/L,
and the range is 0.2 to 150 ^g P/L. Chamberlain and Shapiro (1969) also
developed a method based on extracting phosphomolybdate into isobutanol, which
is followed by reduction of the organic extract with stannous chloride. Color
development is rapid and stable for a day. The limit of detection is ca.
0.3 |ig P/L, and the range is from 0.3 yg to 180 yg P/L.

In the last 15 years, three additional approaches have been used to
measure SRP at low concentrations in samples where the long acid exposure times

of conventional methods may overestimate true P04--P because of the hydrolysis
of P-containing organic compounds or because PO4-P is released from particles
in filtrates. The first method is that of Rigler (1966), who developed a
bioassay method where -^^P (ortho phosphoric acid) and different amounts of
reagent PO4-P are added to samples of lake water. By measuring and plotting
the uptake velocity of PO4-P against the PO4-P concentrations in samples
(assuming different original concentrations of PO4-P in the water), the maximum
value for the true PO4-P concentration in the sample can be obtained. The
method is time-consuming, laborious, and requires specialized equipment, but it
is very sensitive and highly reproducible. Comparisons of PO4-P concentrations
measured by radioassay and SRP values determined by conventional methods
generally show that SRP is one to two orders of magnitude higher than true
PO4-P concentrations .

The second approach, developed by Crouch and Malmstadt (1967), is based on
the fact that as sample acidity is reduced ,> the rate of color development of
reduced phosphomolybdate increases. By lowering acidity and by measuring
optical density after a constant time inteirval, it is possible to perform a
sensitive determination before substantial amounts of hydrolysis occur. A
similar approach, a 6-sec method, has been developed by Chamberlain and Shapiro
(1969). Acid molybdate and reductant are added and color development is
measured in 6 sec to circumvent hydrolysis of all but the most labile
compounds. This method is highly sensitive and generally gives lower SRP
concentrations than conventional methods. The third approach is a new
enzymatic technique that currently is in a developmental state (Pettersson
1979). It is based on the ability of PO4-P to competitively inhibit the
hydrolytic activity of alkaline phosphatase from Escherichia coli . Data from
Lake Erken, Sweden, show that PO4-P values measured by the enz3rmatic method are
lower b3r an order of magnitude than estimates based on conventional analytical
methods .

INTERFERENCES

A serious error in most conventional molybdenum blue methods, whether used
to measure TP, SP, or SRP, arises from interferences by arsenate or other ions
in the same family as PO4-P. The ions complex with molybdate and are reduced
by stannous chloride giving a blue color, v^hich leads to an overestimate of
PO4-P concentrations. This problem is severe in lakes where arsenicals have
been used to poison algal blooms or rooted aquatics (see Shapiro 1971).
Arsenate interferes in Harvey's method but arsenomolybdate is not extracted
into the organic solvent of extraction methods, and it interferes only slightly
in the ascorbic acid method when color development time is kept below 12 min.
Silica also can interfere in PO4-P analyses by forming silicomolybdate which is
reduced by stannous chloride or other reductants. Silica interference will
occur, however, only in the range of pH 3.0-4.0, which is much higher than that
of most colorimetric methods where the pH of acidified samples is generally 1.0
or less.

Other types of interferences or errors can arise from "salt effects" or
from "hidden blanks." Salt effects are encountered when the rates of color
development of reduced phosphomolybdate are different in samples of lake water
and in standards prepared in distilled water. P04'"P concentrations can be
underestimated by 20-25% by Harvey's method due to the "salt effect," but the

ascorbic acid method does not appear to be affected appreciably. In all cases,
however, it is prudent to use internal standards, i.e. PO4-P standards prepared
in lake water. Serious errors in all colorimetric methods also can arise from
the "hidden blank" effect or from the non-specific reduction of molybdate
(see Chamberlain and Shapiro 1973). The magnitude of this error, however, is
difficult to assess.

HYDROLYSIS AND DESORPTION

One of the principal problems in SRP analyses in natural waters is that
dissolved organic compounds or particles <0.50 ym in filtered samples may
release PO4-P under acidic conditions, thus leading to significant
overestimates of true PO4-P concentrations. Despite suspicions and
circumstantial evidence for the phenomenon, as well as arguments that
hydrolysis does not occur, relatively little research has been done to identify
the source(s) of PO4-P released into solution under the acid conditions of
molybdenum blue analyses or to assess the magnitude of the release. Rigler
(1966) has shown that colorimetric methods overestimate true PO4-P
concentrations in lake water by 10-100 times, and Kuenzler and Ketchum (1962)
and Lean and Nalewajko (1976) showed that the same phenomenon occurs in
laboratory algal cultures. Similar results using Rigler' s bioassay method have
been generated for offshore Lake Michigan water (Tarapchak et al., 1980a) and
for an oligotrophic Lake in Ontario, Canada (Levine 1975). The fact that
hydrolysis of organic compounds or desorption of PO4-P from particles occurs
during conventional SRP analyses is also supported in theory by the enzymatic
method of Pettersson (1979).

The source of the liberated PO4-P is uncertain, but it has been thought to
be from highly labile organic compounds since the "dissolved organic P pool" in
lake water is considered to be relatively large (see Rigler 1968). It is
unlikely, however, that such compounds are long-lived in the epilimnion of
typical north-temperate lakes, where bacteria would quickly degrade them.
Chamberlain and Shapiro (1973) also argue against dissolved organic compounds
as a significant P source since the average lability of carbohydrate esters
expected to occur in lake water is very low.

The most likely source of much of the molybdate reactive PO4-P pool is
from small particles, probably colloids or other particulates <0.45 ym, that
pass through membrane filters into filtrates. Evidence for this has come from
diverse reports. Tarapchak et al. (1980b) demonstrated that true PO4-P
concentrations could be overestimated by 50-100% using the Chamberlain-Shapiro
(1969) extraction technique and that release of PO4-P from particles <0.45 um
was a significant source. PO4-P associated with a high molecular weight
molecule in water from Lake Tutaeinanga, New Zealand, is liberated by the 6-sec
method of Chamberlain and Shapiro (1969), Paerl and Downes (1978), and
Schindler ejt al. (1972) also found that acid molybdate may hydrolyze PO4-P from
relatively large molecular weight colloids reported by Lean (1973a, 1973b) for
lake water. Stainton (1980) also has clearly demonstrated that molybdenum blue
methods release PO4-P from colloids in filtered water from a Precambrian Shield
lake in Ontario, Canada.

A concerted effort to identify the chemical composition of P-containing
particles in lake-water filtrates should be undertaken. Profitable results
could be obtained by using techniques such as Sephadex gel filtration, ultra

centrifugation, ion exchange resins, or extraction into different solvents
followed by conventional SRP analysis .

FILTRATION ERRORS AND OTHER ARTIFACTS

Serious errors in SRP analyses of lake water can occur when samples are
filtered through membrane or other types of filters and from "bottle effects."
Dissolved PO4-P or small particles with adsorbed PO4-P can be released into
filtrates during filtration by leakage from algae or by breakage of other
particles, or small P-containing particles that liberate PO4-P when filtrates
are acidified can be trapped in the pores of filters by clogging which excludes
them from filtrates (see Rigler 1973). Storage of samples in various types of
containers also can lead to increases or decreases in SRP concentrations.
Increases can occur if enzymes liberate PO4-P, and decreases can occur if
bacteria attached to the walls of the container assimilate dissolved PO4-P. In
order to minimize these problems, very low vacuum pressures should be used,
constant-volume samples should be filtered, and analyses should be performed as
quickly as possible after samples are collected and filtered.

BIOLOGICALLY AVAILABLE PHOSPHORUS

As attempts are made to understand more precisely the relationships
between phytoplankton growth and phosphorus availability in P-limited lakes,
analytical measurements of dissolved PO4-P have come under close scrutiny.
Evidence from some investigations indicates that SRP measurements may not be
equivalent to that fraction of dissolved P that is biologically available to
phytoplankton and bacteria. Two bioassay techniques have been used:
short-term and long-term. Chamberlain and Shapiro (1969, 1973) compared the
results of bioassays using P-starved Microcystis aeroginosa (which gave esti-
mates of biologically available P) with estimates of SRP measured by three
different analytical methods. After corrections for arsenate interference, the
chemical methods gave similar values and they were essentially equivalent to
biologically available P. Walton and Lee (1972) also showed in longer-term
incubations, where growth instead of nutrient uptake was measured, that SRP
measured by the ascorbic acid method agreed with the results of bioassay mea-
surements. In subsequent studies on the biological availability of P in
tributaries entering Lake Ontario and in water from Lake Ontario itself, how-
ever, Cowen and Lee (1976) showed that on the average biologically available P
was equal to the SRP concentration plus 20% of the total phosphorus value of
the sample. These results indicate that a significant fraction of the P pool
that is not measured as SRP becomes available for phytoplankton growth over
time.

These types of results, coupled with the fact that conventional methods
measure SRP rather than truly dissolved PO4-P (see Rigler 1966 and others
above),, are paradoxical. Since algae take up truly dissolved PO4-P (Kuhl
1974), and colorimetric methods overestimate this fraction of the dissolved P
pool in lake water (Rigler 1966), how do algae take up SRP that is not present
as truly dissolved PO4-P in lake water? The most logical explanation for this
phenomenon is that algae do assimilate only truly dissolved PO4-P and that much
of the SRP is made available to the algae by one or both of two mechanisms:

(1) an inducible enzyme, alkaline phosphatase, synthesized by P-starved algae,
liberates PO4-P from dissolved organic compounds (see Herman 1970), or (2) when
the concentration of truly dissolved PO4-P is depleted by algal uptake, the
equilibrium between the dissolved and the particulate PO4-P pool in lake water
shifts, releasing P04~P from particles into solution. Further research is
necessary to identify the chemical forms of P compounds and complexes in the
SRP pool that are biologically available and how rapidly they are produced and
taken up by algae. Information on the cycling rates of different chemical
fractions of P in lake water and factors that may affect their biological
availability are presented in the next chapter.

RE.FERENCES

A.P.H.A. 1976. Standard methods for the examination of water and wastewater.
Amer. Public Health Assoc, Washington, D.C. 14th Edition.

Atkins, W. R. G. 1923. The phosphate content of fresh and saltwater in its
relationship to the growth of the algal plankton. J. Mar . Biol . Assoc . ^
U.K. 13: 119-150.

Berman, T. 1970. Alkaline phosphatases and phosphorus availability in
Lake Kinneret . Limnol . Oceanogr . 15: 663-674.

Chamberlain, W. M. , and J. Shapiro. 1969. On the biological significance
of phosphate analysis; comparison of standard and new methods with a
bioassay. Limnol . Oceanogr . 14: 921-927.

and . 1973. Phosphate measurements in

natural waters — A critique, pp. 355-366. In : E. J. Griffith,
A. Beeton, J. M. Spencer, and D. T. Mitchell (eds.) Environmental
phosphorus handbook . John Wiley and Sons, Inc., New York.

Cowen, W. F., and G. F. Lee. 1976. Algal nutrient availability and

limitation in Lake Ontario during IFYGL. Part 1. Available phosphorus in
urban runoff and Lake Ontario tributary waters. U.S. Environmental
Protection Agency Report EPA-600/3-76-094a. 218 pp.

Crouch, S. R., and H. V. Malmstadt. 1967. An automatic reaction rate method
for determination of phosphate. Anal . Chem . 39: 1090-1093.

Deniges, G. 1920. Reaction de coloration extremement sensible des phosphates
et des arseniates. £.R. Acad . Sci . , Paris 171: 802-804.

Golterman, H. L., and R. S. Clymo. 1971. Methods for chemical analysis of

fresh waters. IBP Handbook No. 8, Third Revised Printing. Blackwell Sci.
Publ., Oxford.

Harvey, H. W. 1948. The estimation of phosphate and of total phosphorus in
sea waters. J[. Mar . Biol . Assoc . U,«K. 27: 337-359.

Harwood, J. E., and W. H. J. Hattingh. 1973. Colorimetric methods of

analysis of phosphorus at low concentrations in water, pp. 289-301.

In ; E. J. Griffith, A. Beeton, J. M. Spencer, and D. T. Mitchell (eds.).

Environmental phosphorus handbook . John Wiley and Sons, Inc., New York.

Kimerle, R. A., and W. Rorie. 1973. Low-level phosphorus detection methods,
pp. 367-379. Iti: E. J. Griffith, A. Beeton, J. M. Spencer, and
D. T. Mitchell (eds.). Environmental phosphorus handbook . John Wiley and
Sons, Inc., New York.

Kuenzler, E. J., and B. H. Ketchum. 1962. Rate of phosphorus uptake by
Phaeodactylum tricornutum. Biol. Bull. 123: 134-145.

Kuhl, A. 1974. Phosphorus, pp. 636-654. In: W. D. P. Stewart (ed.),

Algal physiology and biochemistry. Bot . Monographs , Vol. 10, Univ. of
Calif. Press.

Lean, D. R. S. 1973a. Phosphorus dynamics in lake water. Science
179: 678-680.

1973b. Phosphorus movement between biological important

compartments. J[. Fish . Res . Board Can . 30: 1525-1536.

and C. Nalewajko. 1976. Phosphate exchange and organic

phosphorus excretion by freshwater algae. J[. Fish . Res . Board Can .
33: 1312-1323.

Levine, S. N. 1975. A preliminary investigation of orthophosphate concentra-
tion and the uptake of orthophosphate by seston in two Canadian Shield
lakes. M.S. Thesis, University of Winnipeg, Winnipeg, Manitoba.

Murphy, J., and J. P. Riley. 1958. A single-solution method for the

determination of soluble phosphate in sea water. J^. Mar . Biol . Assoc .
U.K. 37: 9-14.

and . 1962. A modified single solution method

for the determination of phosphate in natural waters. Anal. Chim. Acta
27: 31-36.

Olsen, S. 1967. Recent trends in the determination of orthophosphate in

water, pp. 63-105. In: H. L. Golterman and R. S. Clymo (eds.). Chemical
environment in the aquatic habitat . Proceedings of an I.B.P.-sjnnposium
held in Amsterdam and Nieuwersluis, October 10-16, 1966.

Osmond, F. 1887. Sur une reaction pouvant servir en dosage colorimetrique
du phosphore dans les Pontes, les Aliers. Bull . Soc . Chem. , Paris
47: 745-748.

Paerl, H. W., and M. T. Downes . 1978. Biological availability of low versus
high molecular weight reactive phosphorus. J_. Fish . Res . Board Can .
35: 1639-1643.

Pettersson, K. 1979. Enzymatic determination of orthophosphate in natural
waters. Internat . Revue Hydrobiol . (in press).

Rigler, F. H. 1966. Radiobiological analysis of inorganic phosphorus in
lakewater. Verb . Internat . Verein . Limnol. 16: 465-470.

1968. Further observations inconsistent with the hypothesis
that molybdenum blue method measures orthophosphate in lakewater. Limnol .
Oceanogr . 13: 7-13.

10

1973. A dynamic view of the phosphorus cycle in lakes,

pp. 539-572. In ; E. J. Griffith, A. Beeton, J. M. Spencer, and

D. T. Mitchell (eds.), Environmental phosphorus handbook . John Wiley and

Sons, Inc., New York.

Schindler, D. , D. R. S. Lean, and E. J. Fee. 1972. Nutrient cycling in

freshwater ecosystems. Proc, IBP Symp. on World Ecosystems, Seattle.

Shapiro, J. 1971. Arsenic and phosphate: measured by various techniques.
Science 1971: 347.

Stainton, M. P. 1980. Errors in molybdenum blue methods for determining

ortho phosphorus in freshwater. Can . J. Fish . Aquat . Sci . 37: 472-478.

Stephens, K. 1963. Determination of low phosphate concentrations in lake
and marine waters. Limnol . Qceanogr . 8: 361-362.

Strickland, J. D. H. , and T. R. Parsons. 1965. A manual of seawater

analysis (with special reference to the more common micronutrients and to
particulate organic material). Fish. Res. Board Can. Bull. No. 125 (2nd
ed., revised), Ottawa.

, and . 1972. A practical handbook of

seawater analysis . Fish. Res. Board Can. Bull. No. 167 (2nd ed.,
revised), Ottawa.

Tarapchak, S. J., and C. Robitschun. 1980a. Comparisons of soluble reactive
phosphorus and ortho phosphorus concentrations in southern Lake Michigan.
Submitted for publication.

Tarapchak, S. J., S. M. Bigelow, and C. Robitschun. 1980b. Overestimation
of orthophosphorus concentrations in surface waters of southern Lake
Michigan: Effects of acid and ammonium molybdate. Submitted for
publication.

Walton, C. P., and G. F. Lee. 1972. A biological evaluation of the

molybdenum blue method for orthophosphate analyses. Verb . Internat .
Verein. Limnol. 18: 676-684.

11

C13APTER II

CHEMICAL FACTORS CONTROLLING
PHOSPHORUS CYCLING IN LAKES

Wayne S. Gardner and Brian J. Eadie

Great Lakes Environmental Research Laboratory

National Oceanic and Atmospheric Administration

Ann Arbor, Mchigan

13

INTRODUCTION

Phosphorus is a principal nutrient limiting primary production in most
fresh waters including the Great Lakes (PLUARG 1978). Phosphorus enters lake
systems through tributaries and land runoff (Schaffner and Oglesby 1978,
Gakstatter and Allum 1978), direct discharge from human activities (Heidtke
_et al . 1979), and atmospheric sources (Eisenreich et al. 1977, DeLumyea and
Petel 1977). An estimate of total phosphorus loads to the Great Lakes is
provided in Table 1 (Chapra and Sonzogni 1979). Diffuse tributary input,
resulting in large part from agricultural land runoff, accounts for a large
portion of total phosphorus added to the Great Lakes (Chapra and Sonzogni
1979). The effect of phosphorus added to or regenerated in a lake depends not
only on the total quantity of phosphorus in the water but also on chemical
form, availability, and residence time in the water. These factors are
initially a function of the phosphorus source but ultimately are controlled by
in situ processes.

Phosphorus chemistry in lakes is simplified by the fact that the element
exists in nature almost completely in the oxidized form as phosphate (Stumm and
Morgan 1970, Williams 1978). However, other factors, including the strong
tendency of phosphate to sorb to and be released from particulate materials,
complicate understanding of phosphorus cycling in aquatic systems. Phosphorus
in lakes can be categorized according to chemical form or classified by size
into dissolved, particulate, and colloidal components. Chemical data on
phosphorus cycling in the Great Lakes have been collated and summarized by
Torrey (1976) and Upchurch (1976). Laboratory studies and field observations
have provided insight into the chemistry and biological availability of
different phosphorus forms in natural waters, but exact mechanisms and rates of
many reactions affecting phosphorus cycling are not known (Torrey 1976). This
chapter will emphasize aspects of phosphorus chemistry which affect nutrient
cycling in the Great Lakes. An attempt will be made to briefly describe known
concepts and indicate areas where future research is needed.

CHEMICAL FORMS OF PHOSPHORUS

Phosphorus in aquatic ecosystems occurs as phosphate in organic and
inorganic compounds. Free orthophosphate is the only form of phosphorus
believed to be utilized directly by phytoplankton (Rigler 1973) and thus
represents a major link between organic and inorganic phosphorus cycling in
lakes. In natural waters, orthophosphate occurs in ionic equilibrium, i.e.

pK =2.2 pK = 7.2 pK - 12.3
as H3P04--< ^H2P04- -^ ^E?0^^^^ ► PO43-

with H2PO4- and HPO42" being the predominant species over pH range of 5 to 9
(Stumm and Morgan 1970). Figure 1 approximates the inorganic species
distribution over pH ranges which occur in the Great Lakes. Complexation of
phosphorus by major ions is significant in sea water; only 56 percent of the
HPO^ and 0.4 percent of the PO43- are estimated to be in free ionic form

15

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16

Fig. 1. Species distribution of H3PO4 in distilled water
(25^0. Data from Gulbransen and Roberson (1973).
Complexation in natural waters alters distribution.

17

(Kester and Pytkowicz 1967). Ca^"*" is the dominant complexing ion in ocean
water (Chughtai _et al . 1968). Although calcium concentrations are lower by an
order of magnitude in Great Lakes water than in sea water, calcium is the major
cation in the Great Lakes and may complex with dissolved phosphorus.

Dissolved inorganic phosphorus also occurs as condensed phosphates which
originate from biochemical phosphorus compounds (e.g. ATP; Holm-Hansen 1969) or
may be derived from detergents in waste waters (Heinke 1969). In addition to
soluble forms, inorganic phosphates may be incorporated into minerals or can be
sorbed to lake sediments (Hwang et al. 1976, Williams ejt £l. 1980), particulate
organic material (Williams ejt al . 1958), clay particles (McCallister and Logan
1978), or inorganic precipitates such as metal hydroxides or oxides or CaC03
(McCallister and Logan 1978, White and Wetzel 1975).

Phosphorus occurs in organisms as orthophosphates, phospholipids, nucleic
acids, nucleoproteins, coenzymes, and phosphorylated sugars (Alexander 1961).
Phosphate is sometimes accumulated by phytoplankton in excess of their
metabolic needs as "luxury uptake" (Reynolds 1978; Lehman, Chapter IV).

In addition to being a cellular component of organisms, organic phosphorus
occurs in natural waters as dissolved organic forms or as non-living
particulate organic compounds. These non-living fractions include residues of
plants, animals, or microbes, and diagenetic breakdovm products of these
residues. Although organic phosphorus compounds originate from incorporation
of phosphate into living material, organic components of aquatic systems may
contain organic phosphates in forms considerably modified from that present in
living material (e.g. humic substances, Nissenbaum 1979). Readily degradable
(labile) organic phosphorus compounds by definition do not exist for long
periods in the water after being released from organisms by excretion or death.
Some organic phosphorus compounds formed by plants and other organisms appear
relatively resistant to decomposition. For example, phytic acid (inositol
hexaphosphate) has been found to be resistant to decomposition in natural water
(Rodel e£ al. 1977). Other compounds (nucleotides and sodium glycerophosphate)
have intermediate stability and decompose over a period of weeks in laboratory
experiments (Rodel et al. 1977).

Of all phosphorus components in natural waters and sediments, the
non-living organic fractions are probably the least understood and may be
chemically the most complex. A primary reason for this lack of understanding
is the absence of suitable standard methodology to isolate and measure
component compounds. Most analytical schemes commonly used for phosphorus
measurements in natural waters quantify organic compounds by difference rather
than by direct measurement of component compounds (Strickland and Parsons 1972,
Standard Methods 1976). To achieve a comprehensive picture of organic
phosphorus composition and activity in lakes, specific compounds or groups of
compounds should be isolated and measured. Methods recently developed for
specific groups of compounds have been used to verify the presence of inositol
phosphates (Weimer and Armstrong 1977, Eisenreich and Armstrong 1977, Herbes
£t al. 1974), nucleic acids (Minear 1972), and ATP (Riemann 1979), but the
composition of a large fraction of dissolved and particulate organic phosphorus
materials remains unidentified (Nissenbaum 1979).

Division of organic material into dissolved and particulate fractions is
generally done by filtration (Standard Methods 1976). Materials passing
through a 0.45 micron pore size filter are arbitrarily labeled as dissolved and
those retained by the filter as particulate. A third category, the colloidal
fraction, cannot be strictly termed as dissolved or particulate but can
logically be considered as a third independent component. Colloidal material

18

constitutes a large fraction of organic material in lake water (Allen 1976) and
may be relevant to phosphorus since much dissolved phosphorus in natural waters
is associated with high molecular weight organic material (Downes and Paerl
1978)-

Categorizing phosphorus on the basis of solubility or other operationally
defined characteristics may be misleading because phosphates can rapidly
exchange between different aquatic compartments (Rigler 1956). Isotopic
studies in conjunction with molecular size fractionations have been used to
define interactions between ortho phosphate, particulate phosphorus, and
dissolved organic phosphorus in lake water. For example, phytoplankton have
been shown to release low molecular weight (ca 250) organic phosphorus
compounds which subsequently combine into high molecular weight colloids with
partial loss of orthophosphate back into solution (Lean 1973). Orthophosphate
can be released from organic phosphorus compounds in water by enzymes (e.g.
phosphatases) or from colloidal humic material by low doses of ultraviolet
light (Franco and Heath 1979).

CHEMICAL REACTIONS AB'FECTING PHOSPHORUS CYCLING

Reactions that can affect phosphorus cycling in lake water and sediments
include precipitation (and f loculation) , complexation, redox reactions,
sorption, acid-base, and a variety of biochemical reactions (Lee and Hoadley
1967). These reactions are in turn affected to varying degrees by factors such
as pH, Eh, temperature, and light. Thermodynamic considerations allow
prediction of forms and relative concentrations of inorganic phosphorus under
equilibrium conditions when Eh and pH conditions and chemical composition are
known. Potential roles of precipitation, complexation, and redox reactions and
maximum levels of inorganic components can be predicted thermodynamically using
solubility or complexation products when chemical conditions of a parcel of
water are defined (Stumm and Morgan 1970). Equilibrium models such as WATEQ
(Truesdell and Jones 1973) can be used to estimate thermodynamic inorganic
phosphate partitioning with major cations and anions in the Great Lakes under
various winter and summer conditions (Table 2). In this model, HP04^'' is the
major form of inorganic phosphorus in the water followed in importance by
calcium and magnesium complexes of phosphorus.

Thermodynamic calculations based on laboratory observations should be
extrapolated to lake waters only with caution because of the many complex
features which make lake systems different from conditions in the laboratory.
A major factor complicating phosphorus chemistry in lakes is the presence of
organisms which dramatically affect nutrient cycling and limit the usefulness
of thermodynamic calculations to predict composition and concentrations of
nutrient forms (Lee and Hoadley 1967).

Complexation and Precipitation Reactions

As mentioned above, phosphate ions can form soluble complexes or insoluble
precipitates with cations, such as Fe, Ca, Mg, and Al, if concentrations of the
metals and phosphates are sufficiently high. Formation of soluble phosphate
complexes in lake water may not affect availability of soluble phosphorus to
organisms but may prevent precipitation or sorption of phosphates when

19

TABLE 2. Equilibrium distribution of phosphorus species.

Lake Michigan

Summer Winter

Lake Ontario

Summer Winter

Lake Superior

Winter

Input
Temp ^C
pH

Ca mg/I
Mg
Na
K
CI
SO 4
HCO3
Si02
Fe
PO4

Output

FeaP04'^

FeHP04

FeR2P04"^

FeH2P04"^

PO4

tlP04 =

Il2P0^~

MgP04"'

MgHP04

MgH2P04''

CaP04

CaHP04

CaH2P04"^

KIIPO4O

NaHP04

20

2

20

2

8.5

3.3

8.3

8.1

34.

34.

42.

42.

11

11.

8.

8.

5

5.

13.5

13.5

1.1

1.1

1.4

1.4

7.0

7.0

30

30.

16

16.

29

29.

128.7

128.7

114.

114.0

I.O

1.0

0.3

0.3

0.01

0.01

0.05

0.05

0.01

0.01

0.01

O.OI

As % of inorganic P

0.01

0.05

0.01

0.008

0.11

0.03

58.

66.7

57.8

67.1

2.4

3.0

2.4

8.6

7.0

3.9

4.9

1.2

9.4

7.9

7.0

6.1

0.02

0.01

0.01

0.04

9.4

6.0

LI. 9

2.8

13.3

I ..2

16.1

U.I

0.04

0.04

0.04

0.10

2

8.1
12.9

2.7

1.3

0.5

1.2

2.3
51.4

3.0

0.1

0.001

0.13

0.003
75.6
9.3
0.6
2.8

L. L
6.1
4.4

0.02

0.02

0. C^

0.08

20

phosphorus levels are high. Under conditions of sufficient phosphate and
calcium levels and high pH, calcite may be converted to hydroxyapetite (Stumm
and Morgan 1970). A review of solubility products and expected levels of
phosphates and metals in sediment waters led Syers ejt al. (1973) to conclude
that metal precipitates of phosphorus [ALPO4 (variscite) , FeP04 (strengite),
and Cay)(P04)5(0H)2 (hydroxyapatite)] are not important for removing phosphate
from sediment water solution. An exception may be hydroxyapetite in some
lakes, but amounts of this precipitate in lake sediments are low (<5% of
inorganic phosphorus; Schofield 1968) and do not appear to be quantitatively
important (Syers et al. 1973).

Sorption, Redox, and Acid-base Reactions

Phosphates are strongly sorbed by a variety of particulate materials
including clays, organic particulates, and various precipitates of metals.
Examples of the latter are the common use of CaC03 and alumina to remove
phosphorus from waste water in treatment plants. Phosphorus does not
participate directly in redox reactions under most environmental conditions but
can be sorbed by precipitates which form as a result of redox conditions. In
natural waters containing Fe, Mn, or Al, orthophosphate is sorbed to the oxide
and/or hydroxide gels of these metals (which form under oxygenated conditions)
but is released when the precipitates dissolve under reducing conditions (Syers
et al. 1973). Removal of phosphate from solution by this mechanism in
oxygenated waters and release of phosphorus back into reduced waters is a
well-known phenomenon in lakes and may be the most important chemical
phenomenon affecting phosphorus distribution and cycling in aquatic systems.
The quantity of phosphorus sorbed to metal precipitates can be reduced by
organic anions which chelate the metals (Earl e^ al . 1979).

Humic substances in sediments can bind phosphates from overlying lake
waters . Nutrients added to hypolimnetic waters of a small Canadian
experimental lake caused minimal eutrophication effects, relative to effects
caused by addition of nutrients to surface waters of a similar lake (Schindler
eX al. 1980). Phosphorus added to the hypolimnion was transferred efficiently
to colloidal material in the sediments with no detectable return to overlying
waters. Phosphorus dynamics in the Great Lakes may be different, however,
because Great Lakes sediments likely contain less humic material and have
higher redox potentials than do the experimental lakes. Great Lakes waters are
subject to large scale physical processes (Boyce 1974) which may result in more
nutrient exchange than occurs in the Canadian experimental lakes.

Three of the Great Lakes (Michigan, Erie, and Ontario) undergo a seasonal
epilimnetic autogenic precipitation of CaC03 (Strong and Eadie 1978). These
particles are postulated to serve as efficient substrates for phosphorus
sorption. Their large size (ca 10 ym) and dense structure cause them to sink
rapidly and remove phosphorus from euphotic zones. After reaching the
hypolimnion the particles in part dissolve and thereby release transported
phosphorus back into solution (unpublished data).

Phosphorus can attach to clays by chemically bonding to the positively
charged edges of the clay particles or by substituting phosphates for silicates
in the clay structures. Phosphates most effectively sorb to clays at pH's of
less than neutral (Stumm and Morgan 1970). Since clays do not normally
dissolve under reducing conditions, sorption to clays presumably is less

21

dependent on Eh conditions than is sorption to oxidized metal precipitates •
Organic phosphates (e.g. nucleotides, nucleic acids, nucleo proteins , inositol
phosphates, and glycerophosphates) are sorbed by clays and other particles
(e.g. Goring and Bartholomew 1952, Anderson and Arlidge 1962). Sorption of
inositol phosphates by clays is greatest at pH 3 to 4 and increases with the
number of phosphates in the organic molecule (Anderson and Arlidge 1962).
Since phosphate esters are sorbed through phosphate groups, sorption of organic
phosphorus compounds can be expected to resemble that of inorganic phosphate
(Syers _et al . 1973). In addition, organic phosphorus compounds may be sorbed
to hydrophobic organic portions of particles.

Acid-base reactions relate directly to pH which controls the relative
concentrations of ortho phosphate forms in solution (under equilibrium
conditions) and may indirectly affect phosphorus chemistry and cycling by
influencing sorption, precipitation, complexation, redox, and biochemical
reactions (e.g. Lee and Hoadley 1967).

Biochemical Reactions

Due to its nutritive importance in lake ecosystems, available phosphorus
is rapidly removed from solution by phytoplankton in photosynthetic zones
(Rigler 1973). This mechanism reduces (and maintains) the amount of dissolved
ortho phosphate in the epilimnion during summer stratification to levels lower
than normally measured by standard analytical techniques (<1 yg L"^, Rigler
1973, Paerl and Downes 1978). Other organisms exert major effects on
phosphorus cycling by immobilizing phosphorus, by making non-orthophosphate
forms available to phytoplankton, or by phosphorus uptake. Interactions of
phosphorus with aquatic organisms and mechanisms affecting its availability
will be discussed in later sections.

Sediment Phosphorus Chemistry

Sediments are important "buffering mediums" which can influence phosphorus
levels in lake waters (Kuo and Lotse 1974). Since photosynthesis is absent
(except possibly at sediment surfaces), biological processes result in a net
mineralization rather than production of organic matter in sediments. Modified
redox potentials and high concentrations of inorganic ions and of sediment
particles, relative to lake water, make the sediment environment potentially
conducive to many reactions discussed above and make sediments an important
potential source of nutrients to lakes.

Factors affecting exchange rates of phosphates and other nutrients across
sediment-water interfaces include physical turbulence, redox potential,
temperature, pH, relative concentration gradients, biological activity, cation
distribution, and sediment composition (Kamp-Nielsen 1974, Holdren and
Armstrong 1980). In many lake sediments, phosphorus is associated with iron,
and levels of the two elements are directly correlated (Syers ejt al. 1973).
The major factor controlling phosphate exchange in anaerobic sediments is the
phosphate concentration gradient (across the sediment-water interface) which is
controlled by pH and redox potential of the sediments (Kamp-Nielson 1974).
Except for the central basin of Lake Erie and polluted bays and harbors,
sediment-water interfaces in the Great Lakes are generally not in a reduced

22

state (Mortimer 1971). Therefore, the mechanism and extent of phosphorus
exchange of oxygenated sediment layers with the overlying water is of
particular interest for these waters.

Exact mechanisms of phosphorus exchange between lake sediments and water
have not been completely defined but involve several interactive processes.
Actual levels of phosphate or other nutrients measured in hypolimnetic or pore
waters do not provide complete information on their flux to and from sediments
because addition and removal processes may occur simultaneously and result in a
small net concentration in the water. -^^P tracer studies have been used to
graphically portray phosphorus exchange between sediment and water as two
(Pomeroy £t al. 1965) or more (Li et al. 1972) first order reactions.
Laboratory and field studies have shown that phosphorus sorption and release by
sediments is affected by time of exposure (Kuo and Lotse 1974, Hwang et^ al.
1976), phosphate concentration and sediment particle size (Hwang et^ _al . 1976),
redox, temperature, mixing intensity, and bioturbation (Holdren and Armstrong
1980). Freshly adsorbed phosphorus is much more "exchangeable" than native
phosphorus in sediments (Kuo and Lotse 1974). Likewise, phosphates recently
sorbed to particulate materials entering the Great Lakes have more potential
biological significance than phosphorus incorporated into minerals or other
refractory particles.

PHOSPHORUS AVAILABILITY TO AQUATIC PLANTS

To understand the impact of phosphorus (from external or internal sources)
on a lake ecosystem, availability of component phosphorus compounds to aquatic
plants must be understood. Any form of phosphorus present in (or added to) a
lake can be categorized as being immediately available, potentially available,
or likely unavailable. The principal form of phosphorus known to be directly
available to plants is or tho phosphate (Hooper 1973). Potentially available
phosphorus forms include a large number of compounds which can be converted to
orthophosphate and thus become available to phytoplankton. This category
includes inorganic condensed phosphates, biochemical forms of phosphorus, and
other sorbed or complexed forms of phosphorus in the water. Phosphates
chemically incorporated into high molecular weight compounds are partially
available to algae (Paerl and Downes 1978). Potentially available phosphorus
can appropriately be further categorized into forms available over short
periods (hours-days) and forms which become available over long periods
(weeks-months) .

The likely unavailable phosphorus compounds are relatively abundant in
lakes but are probably the least understood of the above categories (Nissenbaum
1979). A logical explanation for their abundance is that they are the residual
compounds left after more labile compounds are removed from the water or
particulate material. Although very few forms of phosphorus in the environment
can be considered as totally unavailable over long periods (years), some forms
(e.g. apetite and other minerals) are likely unavailable during their residence
in lakes (Williams et_ al . 1980, Lee e£ al. 1980). More than half the
phosphorus contributed to the Great Lakes by tributaries has been estimated to
be unavailable for plant growth (Chapra and Sonzogni 1979). Rates of
availability and particle residence times in the water must both be considered
to correctly estimate potential biological use of particulate phosphorus added
to lakes (Verhoff and Heffner 1979).

23

Availability of organic phosphorus is not well understood for particulate
or dissolved forms. At least a portion of reactive phosphorus in high
molecular weight organic compounds in lakes can be made available for algal
uptake, but the time required for utilization is longer than for orthophosphate
(Paerl and Downes 1978). Enzyme activity may be responsible for making organic
phosphorus available to phytoplankton (Herbes £t al . 1974, Paerl and Downes
1978).

Chemical Techniques to Estimate Bioavailability

Various methods to evaluate or estimate availability of particulate
phosphorus to aquatic plants have been recently reviewed (Armstrong et al.
1979, Logan et al. 1979, Williams et al. 1980, Lee et al . 1980). Algal assays
(based on increased algal biomass resulting from particulate phosphorus
utilization) have been developed and standardized (USEPA 1971, 1974). These
methods are effective for estimating short-term potential availability but do
not provide complete information on actual availability in lakes. Some
particulate material may be removed from the euphotic zone before potentially
available phosphorus is released (Williams et al. 1980). Other material may
degrade slowly (e.g. some organic phosphates) or may remain suspended in the
water for a long period (e.g. clay-sized particles) and may ultimately be more
available than predicted (Armstrong ^ £l • 1979, Lee et al. 1980).

Chemical extraction techniques using dilute acid, base, or complexation
solutions or ion exchange resins have been used to approximate biological
availability (Cowen and Lee 1976, Armstrong £t al . 1979, Logan et al. 1979,
Williams jet al . 1980). Chemical methods to assess phosphorus availability
must, of necessity, be empirical. One scheme (Allen and Williams 1978) defines
non-apetite phosphorus as that extracted by citrate-dithionte bicarbonate and
apetite phosphorus as that extracted from sediments with dilute acid (HCL or
H2SO4). Dilute acid removes more phosphorus from particulate material than do
algae ( Selanastrium capricornutum) , presumably because it extracts unavailable
phosphorus from apetite and other minerals. Base and anion exchange
extractions of particulate material agree more closely with bioassays than do
acid extractions (Cowen and Lee 1976). It has been proposed that anion
exchange extraction removes phosphorus which is readily available to organisms,
whereas the NaOH extraction is a better indicator of the maximum amount of
inorganic phosphate which could be made available if phosphorus is removed from
solution (Armstrong et al. 1979). Other workers (Logan et^ al. 1979) consider
phosphorus extracted by NaOH to represent short-term availability and the total
quantity of phosphorus sequentially extracted by NaOH and citrate-dithionate
bicarbonate to represent the total phosphorus available. Based on comparison
of chemical studies with bioassay results (Cowen and Lee 1976), available
phosphate was estimated to be roughly equal to soluble reactive phosphorus plus
20 percent of total particulate phosphorus for Great Lakes tributary waters
(Lee et^ al. 1980). Selanastrium capricornutum bioassays on sedimentary
materials from Lakes Ontario and Erie, tributaries, and eroding bluffs
indicated that available phosphorus cannot be estimated accurately by measuring
total phosphorus but is directly related to quantities of inorganic non-apetite
phosphorus in the particulate matter (Williams et al. 1980). Rate of
availability of total phosphorus in Lake Erie tributary water has been examined
by measuring biological (algal) accumulation of phosphorus from test waters
with time over extended (60-80 days) periods (Verhof f and Hef fner 1979) .

24

Processes Making Phosphorus Available to Aquatic Plants

Phosphorus is made available by processes that convert non-available forms
to or tho phosphate or that transport phosphate from unlighted regions into
photosynthetic zones of lakes. Sediments are known sinks for phosphorus
(Mortimer 1941) and are a major reservoir of phosphates which can potentially
be made available to organisms through chemical, biological, and physical
processes. Particulate material can release phosphorus while settling through
the water column and after reaching the sediments. If water and/or sediments
are oxygenated and sufficient iron is pres€int, a ferric hydroxide gel forms and
sorbs dissolved phosphate from solution and thereby prevents (or hinders)
phosphate from becoming available to organisms in the photosynthetic zone
(Syers et al. 1973).

AJ.though it is commonly assumed that oxygenated sediments are strictly a
phosphorus sink, some studies (e.g. Lee et al. 1977) suggest that phosphates
can be released from oxygenated sediments. Although the rates of release are
slower than for reduced sediments, phosphorus contributions from oxygenated
sediments can potentially be substantial and should be considered in
comprehensive models of phosphorus dynamics. Processes that can make
phosphorus available from sediments include biological, chemical, and physical
phenomema. Biological regeneration of phosphates from sediments by benthic
macro-organisms and microbes appears to be an important mechanism to return
particulate phosphorus to water in dissolved form and is discussed in Chapter
VI. Chemical changes in redox potential cause phosphates to be released when
metal precipitates dissolve causing increased phosphate levels in pore water
and in hypolimnetic waters. Physical mixing of the upper layers of the
sediments results in transport of phosphorus-rich pore water with hypolimnetic
water to photosynthetic zones. Sediment material containing exchangeable
phosphorus may also be mixed with bottom water to form a nepheloid layer
(Chambers and Eadie 1980) which can be a mode of phosphate transport when
transported to photosynthetic zones. Availability of nepheloid phosphorus
should be studied in detail; preliminary results by the NaOH extraction
technique from southeastern Lake Michigan indicate that 20-50% of the total P
in the nepheloid layer becomes available. Quantitatively this is approximately
equal to the total phosphorus pool in the v^ater column (Eadie, B. J.,
unpublished data). After phosphorus is transferred from sediments into
hypolimnetic waters, other physical processes (spring overturn, bottom current
movement, and wind-induced upwelling events) can move potentially available
phosphorus into photosynthetic zones (e.g. Mortimer 1969, Stauffer and Lee
1973). These processes should be thoroughly studied to enhance understanding
of Great Lakes nutrient dynamics .

Modeling studies (Scavia 1979) and other work (cf. Hooper 1973) have
suggested that much of the utilized phosphorus in lakes is made available to
phytoplankton through regeneration processes. It can be reasonably
hypothesized that much of the phosphorus is rapidly returned to the
ortho phosphate (or other available) form through excretion or death of aquatic
organisms. Biochemical processes are likely primarily responsible for
converting organism-derived phosphorus compounds back to orthophosphate.
Zooplankton release orthophosphate as part of their normal metabolism (Marshall

25

and Orr 1955, Rigler 1961, Pomeroy £t al. 1963, Barlow and Bishop 1965).
Autolytic enzymes may convert organic forms of phosphorus to orthophosphate
when organisms die. High energy forms of phosphate in living organisms (e.g.
ATP) are not stable for long periods (minutes) after the organism dies
(Christian e£ al. 1975). Phytoplankton produce extracellular enzymes
(phosphatases) which can free orthophosphate from organic phosphorus compounds
(Paerl and Downes 1978). Stability of dissolved organic compounds to enzymatic
decomposition could be a major factor regulating levels of organic phosphorus
in lake water .

Factors Making Phosphorus Unavailable to Organisms

Available phosphorus in lake systems can be converted into forms which are
unavailable to phytoplankton by 1) physical-chemical processes which remove
phosphorus from solution and/ or transfer it from photosynthetic zones and 2)
biological processes which convert utilized phosphorus into refractory
(unavailable) dissolved or particulate forms. Particulate material which does
not degrade will ultimately be transported into the sediments by settling, or
out of the lake with outflowing water. Refractory phosphorus in dissolved or
colloidal organic compounds may become incorporated into particulate material
by forming aggregates or by sorbing to particulates such as clays, inorganic
precipitates, or organic particles. Remaining dissolved refractory material
will move with the water conservatively. The relative quantity of organic
phosphorus in lakes which is refractory over the time scale of months or
seasons is not known, primarily because the composition and relative stability
of component compounds are not known. Also unknown are the relative functions
of plants, animals, and microbes in forming refractory compounds (Hooper 1973).
If turnover is not rapid for some organic materials, phosphorus in these
substances would be measured in the non-living fraction but could become
available later for phytoplankton use. Substances which are rapidly decomposed
would not be present in measurable quantities except for catastrophic occasions
(e.g. death of an algae bloom; Gardner and Lee 1975) which add large amounts of
these compounds over short time periods. Organic forms of phosphorus thought
to be at least partially unavailable over long (<months) periods include some
inositol phosphates (e.g. phytic acid; Eisenreich and Armstrong 1977) and high
molecular weight humic materials (Nissenbaum 1979). Speciation of phosphorus
components of humic and fulvic materials in lakes and their ability to exchange
phosphorus have not been carefully investigated (Nissenbaum 1979).

SUMMARY AND RECOMMENDATIONS

Summary

The importance of component phosphorus compounds to a phosphorus-limited
lake system is determined by their concentration, chemical form, availability
to organisms, and residence time in photosynthetic zones. Phosphorous occurs
(as phosphate) in dissolved, particulate, and colloidal components of lake
ecosystems. Dissolved forms include orthophosphate, condensed inorganic
phosphates, and dissolved organic phosphates. Particulate material can
arbitrarily be divided into living and non-living forms containing 1) naturally

26

incorporated phosphates and 2) sorbed and/or loosely bound phosphates.
Colloidal inorganic or organic phosphates may be important to phosphorus
cycling in lakes, but are not well defined. Organic phosphates in lake water
include inositol phosphates, nucleic acids,, nucleotides, and huunic materials;
the composition of other organic phosphorus components is unclear. Since
orthophosphate is the only form known to be available to plants, other forms of
phosphorus presumably must be converted to orthophosphate to affect biological
dynamics .

Phosphate levels in lakes can be affected directly by complexation and
precipitation reactions, but sorption and biochemical reactions appear to be
the most important chemical factors controlling phosphorus dynamics.
Phosphorus does not normally undergo redox changes in nature, but its form and
concentration in water are often controlled by sorption to (or release from)
colloidal or particulate metal hydroxides which form or dissolve as a function
of redox potential. Phosphorus is thus removed from solution under oxygenated
conditions and dissolved into solution under reduced conditions. Clay
particles, CaC03 precipitates, and colloidal or particulate organic materials
affect lake water phosphorus levels by sorption and desorption.

Eh-iring summer stratification, biological uptake by phytoplankton is a
primary factor controlling orthophosphate levels in phosphorus-limited euphotic
zones. Biochemically accumulated phosphate is returned to the water by
excretion (by zooplankton and other animals), autolytic release after death of
organisms, and (possibly) as a result of degradation of non-living organic
materials. Biochemical reactions which result in the formation of phosphorus
compounds resistant to decomposition are potential mechanisms for making
phosphorus unavailable in lake water. The availability of phosphorus in
nonliving organic dissolved and particulate material is not well known, in part
because the composition of phosphorous-containing organic material in lake
water has not been well defined. Biological methods (based on growth of algae
in phosphorus-limited cultures) and chemical extraction methods, which have
been empirically correlated with algal grov/th, have been used to estimate
availability of phosphates associated with particulate materials.

Sediments are important sinks for phosphorus in lake systems, but can also
be a source of orthophosphate to photosynthetic zones if chemical, biological,
and physical conditions liberate orthophosphate (or other available phosphorus
compounds) and transfer it to the epilimnion. Phosphorus chemistry in
sediments and lower hypolimnetic waters differs from that in surface waters
because concentrations of chemicals and particulate materials are higher and
redox potentials are generally lower in these regions than in epilimnetic zones.

Recommendations '

Processes controlling internal phosphorous cycling must be clarified and
quantified to properly determine effects of changing the levels and/or
compositions of phosphorus from external sources. A comprehensive
understanding of phosphorus aquatic chemistry in lakes is needed to accurately
predict biological effects of added phosphorus. To understand phosphorus
dynamics in the Great Lakes, the composition and cycling rates of major
component compounds must be defined. Specific compounds or groups containing
phosphorus should be identified and their stability and flux rates should be
determined. Quantity, composition, and potential availability of phosphorus

27

compounds entering the lakes from aquatic and atmospheric sources should be
evaluated, so that treatment methods can be designed to remove phosphorus
compounds which would have the greatest impact on lake biota. Physical,
biological, and chemical processes responsible for returning potentially
available hypolimnetic and sediment phosphorus to euphotic zones should be
quantified and compared to that from external phosphorus sources. This
information is required to determine relative effects of reducing phosphorus
inputs to the lake system.

28

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31

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34

CHAPTER III

BACTERIAL DYNAMICS AND PHOSPHORUS CYCLING

Russell Moll

and
Mark Brahce

Great Lakes Research Division

The University of Michigan

Ann Arbor, Michigan

35

The cycling of nutrients through the bacterial component of aquatic food
webs is poorly understood. This lack of understanding can be attributed, in
part, to the wide variety of metabolic activities associated with bacteria.
These processes include: decomposition, mineralization, remineralization,
heterotrophic uptake, nutrient release, and nutrient uptake (Paerl 1978, Faust
and Correll 1976, McFeters £t al . 1978, Ramsay 1976). Such broad descriptors
of microbial activity suffer from having vague definitions which further
obscure the true role of bacteria in the aquatic environment. The purpose of
this chapter is to present a review of nutrient cycling in aquatic bacteria and
demonstrate where fundamental deficiencies in information lie. Particular
emphasis will be placed on the cycling of phosphorus in fresh waters.

The confusion of using an imprecise term to describe a particular process
can be partially avoided by using the simple conceptual model of Barsdate
et^ _al . (1974). This model, shown in Figure 1, identifies the important
nutrient pathways of the phosphorus cycle in bacteria. The model need not be
restricted to the cycling of phosphorus as it illustrates the pathways of most
elements important in bacterial metabolism. For the purposes of this chapter,
movement of elements from water to surplus pool via rl will be called nutrient
uptake; this uptake can involve either organic or inorganic compounds.
Movement of elements from bacterially bound pool or surplus pool to water via
r4 and r2 will be called nutrient release, which can also be organic or
inorganic compounds. Grazing of bacteria includes movement from surplus pool
or bound pool to protozoan pool' via r7 and r9. Pathways r8 (release by
grazers) and r6 and r5 (movement from water to the detritus pool) are discussed
in other chapters.

NUTRIENT UPTAKE

Uptake of organic compounds (via pathway rl) by bacteria is one of the
most common measurements made in aquatic microbiology. This process has led to
the development of several techniques where dissolved radioisotope tracers are
"fed" to bacteria and uptake rates measured. Initially begun by Parsons and
Strickland (1962), and refined by Wright and Hobble (1966), this method has
been used extensively in most types of aquatic environments with a large
variety of organic substrates (Allen 1969, Brown e£ al. 1978, Crawford et al.
1974, Hamilton and Preslan 1970, Vaccaro and Jannasch 1966, Berland et al.
1970). The net conclusion from these experiments is that bacteria are capable
of utilizing a variety of organic substrates at relatively low concentrations
(10""° to lO""-*-^ M) . Although uptake has been measured in many aquatic
environments, a consistent pattern has not been found. ~The variability in
reported uptake rates of different organic compounds indicates that any one
estimate of uptake rate may not be representative of a large number of aquatic
environments; the heterogeneous distribution of aquatic bacteria requires more
reliance on empirical results rather than on estimated rates (Jones 1977).

The limited usefulness of the measurement of bacterial uptake of organic
compounds stems, in part, from problems with techniques. The Wright and Hobble
(1966) method requires the assumptions that uptake of a high concentration
(relative to natural concentrations) of organic substrate is uniform over
several hours (duration of the experiment) and that it follows simple enzyme
kinetics. Contemporary thinking (Williams 1973, Azam and Holm-Hansen 1973)
discounts the possibility these assumptions can be met in a heterogenous

37

Water

r4

r8

r6

r5

Detritus
Phosphorus

r2

r1

Surplus
Bocterioi
Phosphorus

r3

Bound

Bacterial

Phosphorus

r7

Protozoan
Phosphorus

r9

Figure 1. Conceptual model of the
phosphorus cycle through bacteria.
The transfer rates among the important
phosphorus pools are indicated as rl
through r9 in the figure (from Barsdate
et al. 1974.)

38

natural bacterial community. Wright and Hobble's technique was modified to
include a respiration correction (Hobble and Crawford 1969). Most recent
methods favor short incubation periods with substrate concentrations very close
to natural levels (Williams 1973, Azam and Holm-Hansen 1973) and also include a
respiration correction (Dietz and Albright 1978).

The distinction between the new method (Dietz and Albright 1978) and the
Wright and Hobble (1966) method is that in the new method true flux is
estimated while the Wright and Hobble (1966) technique only measures gross
uptake of the substrate at best. The gross uptake is usually reported in terms
of turnover times which have little ecological meaning. Turnover times are
only valid when the rate of substrate uptake does not change throughout the
entire turnover time period. Because many turnover times are several days in
length, it is unlikely that uptake rates remain constant for the turnover time
period (Williams 1973). The new method uses natural levels of substrates and
gives results much closer to true ecological fluxes and also produces more
reproducible results (Azam and Holm-Hansen 1973).

Irrespective of which technique is used, the evolution of the measurement
of nutrient uptake by bacteria will continue. This evolution of the
measurement of organic uptake by bacteria has slowed progress in understanding
the true metabolic activity of aquatic bacteria. Each different technique
provides a different estimate of bacterial nutrient uptake which is not
necessarily comparable to previous estimates. For now, the only conclusion
concerning uptake is that aerobic bacteria do take up a variety of organic
substances at low concentrations in either the dark or light.

The uptake of inorganic compounds by bacteria also has been measured
(Paerl and Lean 1976, Rhee 1972, Phillips 1964). The experimental errors
involved in the measurement of inorganic uptake exceed those with organic
uptake. For instance, inorganic P uptake is often measured as "soluble
phosphate" which is an imprecise chemical descriptor. A major difficulty in
measuring inorganic phosphorus uptake is the rapid rate of uptake and turnover
and competition between bacteria and algae for the same resource; thus
experiments where ^^P or -^-^P are added to natural populations may be difficult
to interpret because of interspecific uptake. Paerl and Lean (1976) and Rhee
(1972) demonstrated relatively rapid movement of inorganic P into bacteria
followed by slower movement into algae. Similar to the uptake of organic
compounds, bacteria are capable of taking up inorganic P compounds at very low
concentrations (Rhee 1972).

Bacteria unquestionably utilize organic substrates for metabolism (ZoBell
1946). These high-energy compounds are the primary source of energy for
heterotrophic organisms. On the other hand, the uptake of inorganic compounds
by bacteria is not fully understood. Presumably bacteria would only take ~up
inorganics to balance their C-N-P ratios. Yet the uptake of inorganics by
those organisms which are supposedly "decomposers" seems inconsistent. Would
bacteria attempt to take up the same compounds they are releasing back to the
water? This paradox, analogous to the extracellular release mechanism in
phytoplankton, will require careful analysis before satisfactory conclusions
can be reached.

39

NUTRIENT RELEASE

The release of nutrients by bacteria (via pathways r2 and r4) in either
organic or inorganic form is the essence of microbial decomposition. Yet this
very important process has been difficult to quantify. Bacterial activity is
influenced by temperature, availability of substrate, and chemical composition
of the substrate (DePinto and Verhoff 1977, Foree and Barrow 1970). Jones
(1972a) demonstrated the relationship between senescent algae and bacterial
biomass; a rapid increase in the amount of decaying algal cells in the water
column, such as when a phytoplankton bloom declines, is usually followed by an
increase in bacterial biomass.

Despite large variability in decomposition rates, some preliminary esti-
mates of inorganic release by bacterial metabolism have been made. Field
estimates of P turnover have been particularly inaccurate in that liberated P
is rapidly re-used and rarely measured as available soluble P. Fuhs (1973) and
Charlton (1975) estimated that water column P was recycled from" 10 to 34 times
during one growing season. Fallon and Brock (1979, 1980) have estimated that
60-70% of the organic matter from decaying phytoplankton was decomposed in the
water column, primarily in the epilimnion. These studies indicated that very
little phosphorus was returned to the water column from anaerobic benthic
release, as these lakes were highly stratified and had little vertical mixing.
Evidently, this active recycling of P in the water column was a consequence of
zooplankton excretion and active release from bacterial decomposition of
detritus .

More exact rates of P release by bacteria are available from laboratory
studies. Golterman (1973) found that 70-80% of the organic P released by algal
autolysis was converted to inorganic forms within a few days. Depinto and
Verhoff (1977) conducted a study of the regeneration of P and N in axenic algal
cultures and bacteria-algae cultures. They found the P released as inorganic
soluble reactive phosphorus was very slow when the axenic cultures were placed
in the dark. On the other hand, the bacteria-algae mixtures liberated up to
75% of the total P pool after several weeks. The rate of regeneration of the
soluble P was a function of the initial P composition of the algae; P-limited
algae produced regeneration rates of 0.06-0.08 yg P/mg algae (dry weight) /day,
while algae growing in P-rich media produced regeneration rates of 0.16-0.39 yg
P/mg algae (dry weight)/day. DePinto and Verhoff (1977) stated that excess
cellular P is stored in algae in an inorganic state. This would explain the
much higher P regeneration rates from algae in a P-rich environment.
Therefore, bacterial crops associated with large P-rich algae communities would
likely be adapted to high levels of inorganic P while bacteria associated with
P-limited algae would have slower regeneration rates and use more organic
compounds for their P source.

The release of nutrients under anaerobic and aerobic conditions is
entirely different. In the absence of oxygen, only reduced compounds such as
methane or hydrogen sulfide can be released (Lerman 1978). Anaerobic
metabolism is extremely important in the sediments and at the sediment-water
interface of some lakes (Wetzel 1975). But, for most lakes and the oceans the
entire water column is oxygenated, confining anaerobic nutrient release to a
small portion of the water body.

Although the rate of nutrient regeneration by bacteria is relatively
poorly quantified, some effort has gone into incorporating this term (or
function) into aquatic ecosystem models. Initially, the simple first-order

40

rate coefficient of "phytoplankton decay,*' usually as a function of
temperature, was used in the models (Thomann et al» 1975). Additional models
required the refinement of this term (as with Lehman jet al. 1975), until
bacterial activity has been treated as a complete sub-model (Clesceri et al>
1977, Bloomfield 1975). The modeling of F cycling in aquatic ecosystems is
discussed in more detail in Chapter VIII. The state-of-the-art in aquatic
models indicates that bacterial regeneration of inorganic nutrients is
apparently a combination of substrate form and availability, water temperature,
bacterial biomass, and bacterial composition (DePinto 1979). Extensive
laboratory and field analysis will be needed before more precise models can be
developed.

NUTRIENT REDISTRIBUTION

The advent of the epifluorescent method for counting bacteria which are
either attached or free-living (bacterioplankton) dramatically changed the
concept of where most bacteria occur. Early dogma suggested bacteria were
primarily, if not exclusively, "decomposers" attached to detritus. But, the
epifluorescent techniques developed by Bell and Dutka (1972), Zimmerman and
Meyer-Reil (1974), and Hobble £t al« (1977) indicated that the vast majority of
bacteria are bacterioplankton (Palumbo and Ferguson 1978). No doubt numerous
bacteria are found on almost every piece of aquatic detritus including decaying
algae (Paerl 1978) and fecal pellets (Ferrante and Ptak 1978). These attached
bacteria could be considered the classic "decomposers," while bacterioplankton
behave more like phytoplankton.

Indirect evidence suggests that bacterioplankton can alter nutrient
distributions (Lean 1973, Jones 1976). Mthough bacteria take up primarily
organic compounds while phytoplankton take up inorganic forms (Lean 1973),
bacteria have been observed to readily take up inorganic P at rapid rates
(Paerl and Lean 1976). Thus, bacterioplankton can effectively compete with
phytoplankton for available soluble P. One major difference is that the
recycle times for bacteria are on the order of minutes, while phytoplankton are
capable of luxury consumption and internal storage (Rhee 1972).

The trophic dynamics of bacterioplankton and phytoplankton are slowly
beginning to emerge. Jones (1971, 1972a, 1972b, 1973, 1976, 1977) has examined
the interrelationships between bacterioplankton and phytoplankton in
freshwaters. He finds bacterioplankton the ultimate opportunists while
phytoplankton are more strategists. Phytoplankton time their growth to the
ideal combination of nutrients, light, and physical conditions (Hitchcock and
Smayda 1977); once ideal conditions are met, phytoplankton grow until one
factor becomes limiting. Bacterioplankton often use a declining phytoplankton
bloom to initiate their growth (Jones 1977). The increased substrate, both POM
(particulate organic matter) and DOM (dissolved organic matter), presumably
serves to stimulate the bacterial bloom.

These growth strategies have large implications for nutrient cycling in
aquatic environments. Bacterioplankton readily convert large amounts of DOM,
originating from either algal lysis or extracellular release, into POM (as
bacteria). DePinto (1979) discussed how these nutrients, taken up by the
bacteria, in turn may be released to the water column as readily available
inorganic nutrients. Thus, as an algal bloom dies off, the bacterioplankton
consume a large amount of the available nutrients and maintain them in the

41

upper water column. Those bacteria which attack decaying algae can have an
impact on the composition of the detritus settling out of the water column.
For example, if bacteria do not take up every element (e.g. silicon), some will
be recycled at very slow rates, while others (e.g. P, N, or C) will be
regenerated much faster.

GRAZING OF BACTERIA

An important pathway (r7 and r9) for the movement of nutrients through the
aquatic environment is grazing of bacteria (Barsdate ejt ad. 1974).
Bacterioplankton are rapidly grazed by small ciliated protozoans, which in turn
are readily consumed by small zooplankton (e.g. rotifers). The ecological
implications of the grazing of bacterioplankton have led Azam (personal
communication) to postulate an entire heterotrophic food web. In this scenario
bacterioplankton readily survive on DOM from the extracellular release and
lysis of dying algae (observed by Jones 1977, Paerl 1978, and others). The
bacterioplankton are grazed by small ciliated protozoans (Barsdate et^ al .
1974). These protozoans serve as ideal food sources for rotifers (J. Bowers,
personal communication), which in turn are food sources for carnivores. The
recycling of nutrients in this heterotrophic food web is carried out by both
attached bacteria and bacterioplankton.

Barsdate ejt al. (1974) conducted an investigation of phosphorus cycling
between bacteria and the protozoans. Using freshwater bacteria, they found
relatively little inorganic P release, but substantial organic P release. The
amount of P regenerated from the bacteria to the water was very sensitive to
the presence of ciliate bacterial grazers. When bacteria were present in
microcosms alone or with small amounts of algae, regeneration rates of
inorganic P were similar to those of P-limited algae-bacteria cultures measured
by DePinto and Verhoff (1977). But, when protozans were added as grazers in
the Barsdate ejt al. (1974) microcosms, the inorganic P regeneration rates were
too low to measure. Thus, according to Barsdate e_t al . (1974) and Fenchel and
Harrison (1976), the concept of bacteria releasing large amounts of inorganic P
as "decomposers" may be incorrect. However, these studies cannot be considered
exhaustive, particularly in that the experiments were conducted in microcosms
and included only one species of protozoan.

The grazing of bacteria attached to detritus has also been suggested as an
important nutrient pathway (Paerl 1978). Detritivores have been observed to
consume large amounts of detritus coated with bacteria (Paerl e^ al* 1975).
Some researchers postulate that detritivores harvest detritus as a convenient
method of consuming bacteria rather than obtaining nutrients from the detritus
itself. The dynamics of grazing of bacteria are not well understood (see
Chapter IV for more information on herbivorous grazing) . While most of the
bacterial grazers have been identified, the type of bacteria commonly grazed,
the percent of the population grazed, and the temporal pattern of the grazers
are unknown .

42

BENTHIC BACTERIAL PROCESSES

The activity of benthic bacteria is probably the least understood of all
bacterial processes. The benthic substrate of aquatic environments provides a
diverse habitat in that anaerobic sediments are in contact with aerobic
sediments and/or the overlying waters. As mentioned earlier, anaerobic and
aerobic bacteria are very different with regard to the compounds they take up
and release. As a consequence of this difference, "decomposition" by the
benthos can provide a broad variety of nutrients to the overlying water.
Benthic bacteria microbial activity has been identified with P release
(Ayyakkannu and Chandramohan 1971). Presumably, benthic bacteria are
responsible for the cycling of all elements which reach the sediments. This
assumption can be made from mass balance calculations, as well as some direct
measurments of near-bottom nutrient fluxes. Fallon and Brock (1980) show that
in Lake Mendota decomposition of blue-green algae occurs in both the water
column and the surface of the sediments; 57% of the decomposition occurs in the
water column, and 32% at the sediment surface. However, a wide variety of
benthic organisms are known to actively "work" the substrate (see Chapter VI).
These organisms must be responsible for some of the near-bottom nutrient flux.
The true activity of benthic bacteria remains in doubt. Jannasch and Wirsen
(1973) claimed that the low temperatures and/or high pressures of the deep sea
inhibit bacterial activity and cause aphotic decomposition or remineralization
to be a slow process. Azam and Holm-Hansen (1974) refuted that claim with data
showing considerable heterotrophic activity in the deep sea.

The actual site of the majority of nutrient regeneration is unresolved.
Benthic bacteria cannot account for all of the recycling. The benthic
microbial biomass is too low in comparison to bacterioplankton biomass on a M^
basis (Holm-Hansen 1969). The primary site of the recycling of some nutrients
(such as P) is most likely the water column, while for other elements it is the
sediments (Kobori and Taga 1979). Fallon and Brock (1980) indicate that very
little (<10%) of the organic matter from phytoplankton remains in the sediment.
Most of the decomposition occurs in the water column, and the material that
does reach the sediment primarily decomposes at the sediment-water interface.

SUMMARY

Progress in understanding the cycling of nutrients through bacteria has
been slowed, in part, by a lack of adequate experimental techniques. Some of
these technical problems stem from the difficulties of separating bacteria from
phytoplankton either with biochemical or mechanical techniques. As long as
this imprecise separation continues, true nutrient uptake and nutrient release
rates will evade researchers. The most recent gains in the state-of-the-art in
ecological aquatic bacteriology have been in enumeration. A considerable
number of new, innovative techniques will be required before satisfactory
progress is made in understanding nutrient cycling through aquatic bacteria.

43

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Bell, J. B., and B. J. Dutka. 1972. Bacterial densities by fluorescent
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Ferrante, J. G. , and D. J. Ptak. 1978. Heterotrophic bacteria associated with
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mineralization of organic phosphorus In coastal sea water. J[. Exp . Mar .
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Important forms In lake water. £. Fish . Res . Board Can . 30: 1525-1536.

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rationales of a computer model of Phytoplankton population dynamics .
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46

Paerl, H. W. 1978. Microbial organic carbon recovery in aquatic ecosystems.
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_, and D. R. S. Lean. 1976. Visual observations of phosphorus

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significance of detritus formation during a diatom bloom in Lake Tahoe,
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particulate organic carbon by heterotrophic processes in sea water.
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plankton in a New Zealand freshwater lake. N. Z. J. Freshwat . Res .
10: 77-90. "~ ""

Rhee, G-Yull. 1972. Competition between an alga and an aquatic bacterium
for phosphate. Limnol . Oceanogr . 17: 505-514.

Thomann, R. V., D. M. Ditoro, R. D. Winfield, and D. J. O'Connor. 1975.
Mathematical modeling of phytoplankto n in Lake Ontario 1_. Model
development and verification . USEPA Report No. EPA-660/3-75-005. 177 pp.

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in seawater based on glucose assimilation. Limnol . Oceanogr . 11: 596-607.

Wetzel, R. G. 1975. Limnology . W. B. Saunders Co., Philadelphia. 743 pp.

Williams, P. J. Le B. 1973. The validity of the application of simple kinetic
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Wright, R. T., and J. E. Hobble. 1966. Use of glucose and acetate by bacteria
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Zimmerman, R. , and L. A. Meyer-Reil. 1974. A new method for fluorescence
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Mass. 240 pp-

47

CHAPTER IV: LOWER FOOD WEB DYNAMICS

CYCLING OF P mom PHYTOPLANKTON,
HERBIVOROUS ZOOPLANKTON, AND THE LAKE WATER

John T. Lehman

Division of Biological Sciences

The University of Michigan

Ann Arbor, Michigan

49

INTRODUCTION

Dissolved phosphorus in lake water includes the immediate source of an
element vital to primary biological production. For that reason substantial
efforts have been undertaken to quantify rates of flux between dissolved pools
and the biota. Particular interest has been focused on reactions involving the
phytoplankton, which are the base of the trophic pyramid in pelagic
communities. A guiding premise of the work reported here is that the
interactions and the kinetics of processes in these systems can be understood
only by perturbing the system in some way and measuring the results. This may
be through massive additions of P to a lake (Einsele 1941; Nelson and Edmondson
1955; Schindler 1975, 19977) or through the use of bioassays or tracer
techniques. The latter methodology was pioneered at the whole lake level by
Hutchinson and Bowen (1950), Hayes e£ al . (1952), and Rigler (1956), from whose
work and through subsequent re-interpretations (Rigler 1973) it became very
clear that exchanges between dissolved and particulate phases occurred on the
scale of minutes. Interest in the rapid removal of tracer-P04 from solution
spawned the Rigler bioassay (Rigler 1964, 1966, 1968) in an effort to quantify
true concentrations of dissolved orthophosphate in lake water, and led
ultimately to the model by Lean (1973) of P-exchange among several functional
pools .

RAPID EXCHANGE FLUXES

Basing their rationales on the demonstration that the most biologically
active pools of dissolved P in lake water are often much lower in concentration
than traditional chemical analyses indicate, one school of research has focused
on the kinetic events that follow the addition of carrier-free radiophosphorus
to lake water or algal cultures (Lean 1976, Lean and Nalewajko 1976, Nalewajko
and Lean 1978). By charting the quantity of radiophosphorus that remains in
solution over a time course that may span only minutes, it is possible to
compute rates of exchange between P in the water and in the algae. Because the
results show that the dissolved tracer P turns over so rapidly in natural
waters primarily as a result of biological mechanisms, there is a tendency to
regard this behavior as the fundamental process controlling the concentration
of phosphate. Whereas this may indeed prove to be the case, it provides no
insight to the flux rates responsible for biological production in lakes. What
is measured in these experiments is isotope equilibrium only, not mass flux
associated with de novo synthesis of biomass. The technique measures only one
side of a two-way flux, and thus the results cannot be included in the same
category with net fluxes resulting from excretion or autolysis of cells.

In fact, the guiding assumption in the studies is one of virtual steady
state for mass influx and efflux, an equilibrium system into which temporary
isotopic disequilibrium is introduced. The measurements essentially quantify
the unidirectional movement of molecules across a cell membrane, even though an
equal mnaber of molecules are moving simultaneously in the opposite direction,
resulting in zero net flux. The isotope technique consequently helps identify
a freely exchangeable pool of P that is associated with living cells, and the
kinetics permit the size of the pool to be estimated. To identify processes
and rates that are responsible for accrual of phytoplankton biomass, net rates
of flux between the water and the cells must be measured. Organisms that

51

divide once per day must have the means to double their cellular P contents
each day also. The issue moves, therefore, to time scales and to nutrient
concentrations for which it is possible to measure rates of influx that
substantially exceed efflux from the cells.

RESPONSES TO ADDED SUBSTRATE

The mechanisms of transport-mediated nutrient uptake and uptake-controlled
cell growth have been reviewed recently by Button (1978). Uptake of P and
incorporation in cellular biomass must be viewed as a several stage process
involving external concentrations of P (pQut) > ^ transport component (T), a
labile internal pool of P (Ppool) > ^^^ structural cellular constituents
(^cell) • Reversible formation of a substrate-transport complex is the inter-
mediate step between equilibration of the internal and external substrate pools:

Pout + T-^TPout*^TPpool— T + Ppool (1)

The internal pool comes into dynamic equilibrium with both Pout ^^^ -l^cell* ^^^
sets the rate of cellular synthesis:

Ppool^^cell (2)

Ppool ^^^ Pcell ^ach consist of components that vary somewhat independently of
each other (Rhee 1973, 1974), but this generalized scheme is an adequate
description of events. This scheme makes clear the point that rates of uptake
of P from solution do not directly control growth rate. When cells are at
steady state with constant ambient concentrations in a chemostat, rates of
uptake can indeed be calculated from growth rates and cell quotas of P (e.g.,
Rhee 1973, 1974), but when external concentrations are changing, that approach
is unsatisfactory (e.g., Burmaster 1978). The cell quota (the amount of
nutrient per cell) changes in response to environmental conditions, and it is
closely linked to growth rates (Droop 1968, 1973, 1974; Fuhs £t al. 1972; Brown
and Harris 1978). The allocation of cellular P among different molecular
fractions including polyphosphates and nucleotide-P has been examined
occasionally (e.g., Rhee 1973, 1974) as has the presence of extractable surplus
P (Fitzgerald and Nelson 1966). The precise nature of the exchangeable pool,
here called Ppool > ^^ ^^^ known precisely, but it constitutes only a small part
of total cellular P (Rigler 1973), and is apparently not the place where
surplus P is stored. Much of the surplus P may be stored as polyphosphates,
some in polyphosphate bodies in cell vacuoles or elsewhere (Harold 1966, Jensen
and Sicko 1974, Stewart 1977, Sicko-Goad and Stoermer 1979).

Droop (1973) avoided the issue of identifying separate cellular nutrient
compartments when he advanced the cell quota as though it were the determinant
of nutrient-limited algal growth and illustrated the empirical validity of the
relationship

y = Pm (1 - V^) (3)

for P-limited growth of algae, y is cellular growth rate (d"^), \i^ is the
theoretical maximum rate, Q is the cell quota (ymol P cell-1), ^nd kq is the
minimal amount of P necessary for cell division. Other authors have discounted

52

the general validity of this approach by demonstrating that for other
nutrients, particularly C, N, and sometimes Si, Eq 3 is a poor description
(Caperon and Meyer 1972, Goldman et al. 1974, Goldman and McCarthy 1978) unless
\x^ is regarded as a purely fictional pararaeter not even remotely similar to the
maximum growth rate that can be estimated for cells in steady state growth.
There is no a priori reason to suspect that P should obey one growth model and
that other potentially limiting nutrients should obey another; the
discrepancies arise because cell quotas are not only determinants of growth
rates, they are products of them, too, and because some nutrients are subject
to much greater variability in the cell than are others (Goldman and McCarthy
1978).

Button (1978) solved the relevant equations to evaluate the kinetics
implicit in Eq 1 for the general case. What emerges is a scheme in which
active transport, molecular diffusion, and product inhibition are all involved
to some extent. Growth rates are controlled by rates of supply of nutrients
from the water only to the point at which intracellular conversion of available
nutrient into structural components becomes the rate-limiting step. Maximal
rates at which P can be transported into the cell may be many times greater
than P can be utilized once it gets in the cell. This means that
concentrations at which growth rates become saturated will be much lower than
concentrations required to saturate the uptake mechanisms (Lehman et al. 1975,
Button 1978, Burmaster 1979).

Net rates of nutrient uptake are usually evaluated by Michaelis-Menten
kinetics :

^ = %'^out/(Kp + Pout) (^)

where u is uptake rate (maximum u = u^) , and Kp is the half saturation constant
for uptake. The model in Eq 1 implies that flux rates may deviate from these
simple kinetics under some circumstances (Button 1978). Brown et al. (1978),
in fact, demonstrated empirically that rates of uptake proceed faster at low
concentrations of PO4 than would be expected by fitting a Michaelis-Menten
curve to data obtained at higher concentrations and extrapolating down to the
low values. The effect does not seem caused by the rapid exchange fluxes
discussed earlier, and may imply the existence of several transport systems in
algal cells, each with a differing affinity for dissolved PO4. The rapid
exchange between Pout ^^^ ^pool ^^es suggest that there could be a
concentration of Pout ^^^ which net uptake is zero, but the affinities of the
transport system are so great that the relevant concentration, if it exists,
would defy conventional methods of chemical analysis.

i^iother source of deviations from simple uptake kinetics is the diffusion
limitation that could occur at low phosphate concentrations (Munk and Riley
1952, Pasciak and Gavis 1974, Gavis 1976). Molecular diffusion in the boundary
region around a cell might not be rapid enough to maintain high flux rates at
the cell surface regardless of the affinity of the transport system. These
effects can be very important when ambient concentrations are low, in which
case sizable concentration gradients needed to maintain rapid diffusive flux
cannot be established. Because some methods indicate that ambient
concentrations of PO4 may be very low (see Rigler 1973), this issue may be
important in lakes .

Although there are reasons to suspect that simple Michaelis-Menten
kinetics do not give an adequate description of uptake rates at low

53

concentrations, the half saturation constant Kp obtained from that kinetic
analysis is still regarded as a good measure of the affinity of the transport
system. Ranges of Kp have been documented for many species (see summaries in
Healey 1973, Lehman et al* 1975), and they raise some important questions. For
instance, in all cases the half saturation constant is much greater than the
concentration usually encountered on average in solution. Does this mean that
the uptake system is always far undersaturated? If so, what is the adaptive
advantage of possessing a system with so much unused capacity? A key
consideration is that the maximum possible rate of uptake of P is usually far
greater than necessary to support the maximum possible rate of cell division.
It follows that luxury uptake of P (uptake and storage of P beyond immediate
cellular needs) could be a common occurrence. In fact, if saturating
concentrations are encountered rarely but dependably, luxury uptake may be a
prime mode of sustenance, as will be discussed later.

NUTRIENT RELEASE TO THE DISSOLVED PHASE

When the nutritional demands of phytoplankton are estimated from primary
production data or from nutrient uptake it has been observed that the cells are
utilizing inorganic nutrients at rates far greater than the substances are
being supplied from external sources or from ambient dissolved pools (Barlow
and Bishop 1965; Ganf and Blazka 1974; Lehman 1978, in press; Richey 1979).
Some of the demand for nutrients must be sated by nutrients cycled within the
water column. Two alternative mechanisms have been proposed to account for the
requisite fluxes. Some authors claim that physiological death of phytoplankton
is such an important event (Jassby and Goldman 1974) that autolysis, possibly
aided by microbial attack on moribund cells, can account for most of the
recycling (Golterman 1973). Whatever its cause, cell death and lysis release
phosphate rapidly to the water, probably because of the labile nature of
phospho-ester bonds, and the autolytic enzymes present in the cells. The rapid
efflux of P from living cells detected during isotope equilibrium experiments
(e.g.. Lean and Nalewajko 1976) probably represents diffusion of molecules from
high concentrations inside the cell membrane.

An alternative source for the nutrients is regeneration from the algae
indirectly, through the grazing activities of herbivorous zooplankton.
Nutrients are regenerated to dissolved pools either by excretion or by the
breakdown of incompletely digested remains of egested algal cells. The
liberated nutrients are very reactive biologically, and they are available not
only to the preferred prey of the herbivores, but also to their competitors.
To illustrate the impact that nutrient release from zooplankton can have on the
phosphorus cycle in lake water, it is useful to review what is known about the
mechanisms and rates of feeding and metabolism.

FEEDING BY ZOOPLANKTON

Empirical investigations have demonstrated that grazing can influence the
composition and abundance of algal communities (Gushing 1963, Haney 1973),
particularly during the late spring and summer. Also well documented is the
fact that the grazers are very selective about the types of particles they
ingest (Arnold 1971, Nival and Nival 1973, Berman and Richman 1974, Richman

54

et al. 1977). Some of the selection may be dictated purely by the mechanical
properties of the filtering apparatus (Nival and Nival 1976, Boyd 1976), but
there may also be a high degree of chemosensation and decision-making involved
(Poulet and Marsot 1978). The favorite procedures for measuring feeding rates
rely either on direct quantification of changes in abundance of the food
particles (e.g., Reeve 1963, Frost 1972), or on incorporation by the animals of
a radioactively labelled food (e.g., Rigler 1961b, Geller 1975). Early workers
believed that filter-feeding zooplankton were completely automatic, in the
sense that each animal would always filter a constant volume of water,
regardless of the concentration or types of particles it contained (Fleming
1939, Harvey 1942). Studies by Ryther (1954) and Marshall and Orr (1955),
however, soon made it clear that at high concentrations of food particles the
rate of ingestion levels off to some maximum value. Subsequent work showed
that the observation held for both field and laboratory populations (Rigler
1961b, Reeve 1963, Burns 1966, Parsons ejt £l. 1967). Several different
formulations were advanced to describe the empirical results mathematically,
but the choice of which equation an author used to represent his data appears
to have been subjective. In the hope of finding a single "best" equation,
Mullin e£ al . (1975a) used Frost's (1972) extensive data on feeding rates of
Calanus pacificus to try to distinguish the best empirical model by statistical
means . They found that the data fit the three different models they examined
almost equally well.

Some authors have taken a less empirical approach to the issue, and have
tried to justify grazing models from basic assumptions about feeding behavior.
Gushing (1959, 1968) and Crowley (1973) derived formulations for grazing
zooplamkton from the same constructs that Rolling (1959) used to erect a scheme
for predation rates among animals in general. An alternative approach advanced
by Lam and Frost (1976) and by Lehman (1976) treated grazing as an optimization
process for the grazer in which effort is expended to gather food, and
nutritional reward results from its digestion. Frost (1975) showed that the
feeding behavior of Calanus is consistent with the optimization hypotheses, but
not with the other models commonly in use. This experimental observation still
awaits confirmation from other species.

Beyond the matter of feeding rates on a single food type and their
mathematical representation is the issue of selectivity exhibited by
zooplankton when they are presented with the array of possible food species in
nature. Many authors (e.g., O'Neill 1969, Bloomfield et al. 1973, Park et al.
1974, Vanderploeg and Scavia 1979a, 1979b) have proposed that simple
coefficients applied to each food source can represent both the mechanical
biases of the food-gathering mechanism (passive selection, in the sense of
Frost 1977), and the more active selection/ rejection process based perhaps on
taste (Poulet and Marsot 1978). The approach is basically sound, but the
models used to represent the behavior are probably invalid for a variety of
reasons. First, the animals can change their selectivity in response to the
nutritional quality and abundance of available prey. Wilson (1973) and Richman
ejt £l . (1977), for instance, showed that some copepods change their selectivity
in response to changes in the available particle spectrum. Second, maximal
feeding rates are dependent on the food species. To some extent this is
because the volume ingested plays a role in regulating feeding rates (McMahon
and Rigler 1965, Geller 1975), and because different food species have
distinctly different carbon- to- volume ratios. Also, there are considerable
differences between species in the degree to which they are compressed inside

55

the gut (Geller 1975),

Once food is ingested by a zooplankter, only a fraction of it is
assimilated and the rest is egested. Rates of assimilation and egestion will
influence not just secondary production in a lake, but rates of nutrient
cycling as well. Strong experimental evidence exists for the claim that
digestion efficiencies are specific to individual food types (Lefevre 1942,
Marshall and Orr 1955, Arnold 1971, Porter 1973). Rates and efficiencies of
assimilation may sometimes be identical for taxonomically dissimilar organisms
like Scenedesmus , a green alga, and Asterionella , a diatom, but very different
for more similar organisms like the green algae Scenedesmus , Stichococcus , and
S t auras t rum (Geller 1975, Lampert 1977b). Some authors have occasionally
reported instances of "superfluous feeding" (Beklemishev 1962) at high food
abundances. That is, much more material is ingested than the animal can
possibly assimilate. Conover (1966a) and Lampert (1977b) disclaimed the
occurrence of superfluous feeding among herbivorous zooplankton, but one should
not conclude that assimilation efficiency is totally independent of food
concentration. Schindler (1968) demonstrated that the ratio of assimilated
energy to ingested energy decreases by approximately a factor of 2 when Daphnia
magna is fed on progressively more concentrated diets. In other words, when
food is very abundant and processing rates are rapid, digestion is less
complete than when food is rare. This corresponds to Geller' s (1975)
observations about variable gut passage times. At low food abundances,
ingested material may be retained inside the gut for more than an hour, but
when external concentrations are high, food is egested within 10 minutes.

Experiments suggest that large differences in assimilation efficiency as a
function of food concentrations are unlikely. Differences of a factor of 2
(Schindler 1968) or less (Lampert 1977b) are probably the full range that occur
in natural populations. The picture is a little different when considering the
fate of materials that are ingested but not assimilated by the animals, for a
very simple reason. The fact that Schindler (1968) reported assimilation
efficiencies between 90% and a little more than 40% means that between 10% and
60% of the ingested material may be returned to the water unassimilated.
Regarded in this way the variability of the egested portion is much larger than
that of the assimilated portion on a relative basis. The nutrients associated
with this egested material are either released to inorganic pools immediately
or are subsequently remineralized by microbial decomposers. Releases of
nutrients caused by inefficient feeding are not confined to egestion. Lampert
(1978) showed that uningested cells which are broken or damaged during
filtering activities of the animal may leach dissolved organic carbon to the
water. It can be assumed that nutrients such as P and N are released as well.
Other investigators have reached the same conclusion (e.g.. Gushing 1955,
Conover 1966b, Martin 1970, O'Conners _et £l . 1976, Bowers and Grossnickle
1978). This mechanism may be even more important among carnivorous cyclopoid
(Brandl and Fernando 1975a, 1975b) and calanoid copepods (Anderson 1970, Bowers
and Warren 1978, Kerfoot 1978, Landry 1978), which leave partially eaten prey
and fragments after an attack. The various fluxes associated with imperfect
feeding efficiencies can be important to nutrient recycling at some times of
the year. Their magnitudes must be added to those of the nutrients that are
excreted to the water during the metabolism of assimilated compounds. Another
consideration is that inorganic nutrient stores in the algae (e.g., luxury
consumed stores of P) are probably not assimilated by the zooplankton, and
these constituents are selectively egested. This means that the nutrient

56

ratios in the egested material may be substantially different from the ratios
that are ingested.

Selective egestion is important, because there has been a long tradition
of estimating rates of metabolic excretion of P from zooplankton on the basis
of measured respiration rates (e.g., Satomi and Pomeroy 1968, Ganf and Blazka
1974, Devol 1979). The published 0:P ratios are subject to considerable
variability, which may be caused by variability in the nutritional condition of
the animals (Ikeda 1977). The requisite information may be more difficult to
interpret than was originally anticipated. Lampert (1975) showed that much of
the carbon assimilated by Daphnia pulex is metabolized almost immediately. He
(Lampert 1977a) suggested that a sizable portion of the metabolic demands of
herbivorous zooplankton is met by the immediate use of newly digested food,
rather than by catabolism of body constituents. This means that the P content
of excreted material may not bear any simple relation to the elemental
composition of the animal tissue. Peters and Rigler (1973) and Peters (1975)
in fact used a radiotracer technique to infer that this was indeed the case for
herbivorous zooplankton .

A better method for determining nutrient release from the zooplankton has
been to measure the nutrients directly. There are problems with interpreting
the results, nonetheless. Measured concentrations represent a balance between
rates of nutrient remineralization by the zooplankton and the simultaneous
uptake of the nutrients by phytoplankton. First filtering the algae from the
water is not satisfactory, because animals without food behave differently from
well-fed ones (Conover and Corner 1968, Takahashi and Ikeda 1975, Mayzaud 1976,
Ikeda 1977, Lampert 1978). Similarly, concentrating the zooplankton to obtain
unambiguous measurements of nutrient excretion can mean that food supplies may
be rapidly exhausted, the animals may be damaged, or their physiology,
including the rates and forms of excreted substances, may change (Mullin et al.
1975b). These points are of interest also in estimating the fractions of total
P that are released to the water in organic and inorganic form because
concentrated plankton collected with nets in the field may release as much as
half of the P in organic form (Pomeroy ejt al. 1963, Satomi and Pomeroy 1968,
Le Borgne 1973), whereas careful laboratory studies with undamaged animals find
that most of the release in short-term assays is PO4 (Rigler 1961a, Butler
et al. 1969, Peters and Lean 1973, Ferrante 1976).

TIGHT BIOTIC COUPLING OF RELEASED NUTRIENTS

Despite potentially rapid rates of nutrient remineralization from a
variety of causes, the concentrations of dissolved P in lake waters usually
remain very low during the seasons of highest productivity. One reason that
released nutrients do not accumulate in the water is they are swiftly removed
by the algae and used to enhance their growth. Cycling fluxes intensify and
less algal biomass is needed to maintain a given level of productivity; new
biomass is generated just as quickly as the old biomass is consumed or lysed
and the nutrients are regenerated. Despite low average concentrations of
dissolved nutrients, algae may sometimes be dividing at rates close to their
physiological maximum, fueled by the rapid biotic cycling of P and N (Goldman
et al. 1979), where the emphasis this time is on net fluxes, not the two-way
fluxes that simple isotope equilibrium experiments show. There are two
mechanisms by which this can happen. On the one hand, uptake rates at low

57

concentrations of P may be much greater than would be expected by extrapolating
down from high concentrations, a possibility that has already been discussed.
On the other hand, the scheme might be precisely the one that Goldman et al.
(1979) envisage, with algae encountering very heterogeneous microenvironments
only micrometers in size, some of which are severely enriched with nutrients
owing to remineralization events. If nutrients are rapidly sequestered by the
cells in these episodic and disjoint patches and then subsequently used for
growth (luxury uptake), mean ambient concentrations would indeed remain low and
yet be perfectly consonant with rapid cellular division rates. If one can
safely assert that zooplankton are dominant agents of nutrient release at some
times or places, then this scheme becomes plausible.

Nutrient release from the zooplankton is ultimately tied to their feeding
rates. Some of the nutrients assimilated by the herbivores and used for growth
and reproduction contribute to fluxes through higher trophic levels, but most
of the assimilated nutrients are returned to solution one way or another during
the season. Activities of the grazers cause selective mortality to the
phytoplankton on the one hand, and fertilization on the other. The effects of
the animals on rates of competitive displacement of algal species due to
differential growth rates are potentially much greater than those caused by
grazing losses alone. Nutrients released to the water as a consequence of the
activities of the animals are available to all the algae, not just to the
preferred prey of the grazers. This means that herbivores can depress the
abundances of some algal species through selective grazing, and enhance
productivity of others by nutrient release.

The best estimates for rates of feeding and of nutrient release have been
obtained from relatively large crustacean zooplankton, particularly cladocera
and calanoid copepods. Smaller herbivores have faster metabolic rates per unit
biomass. This means that rotifers and protozoa, for instance, deserve careful
attention. If the biomass of these small metazoa and protozoa is substantial
in any water mass, their impact could be considerable. Gliwicz (1975, 1976)
reported that rates of primary production in two tropical lakes are closely
correlated to the grazing activities of herbivorous zooplankton, and Lehman
(1978, in press) found that nutrient release by zooplankton can supply a major
share of the nutritional requirements of epilimnetic phytoplankton in a
temperate lake during the summer, an observation corroborated by Devol (1979)
for the same lake by independent means. Rapid cycling of nutrients between
grazers and algae promotes high rates of primary production and cell division
at times when allochthonous inputs and dissolved pools could not support
populations for even one full day.

SUMMARY

Processes responsible for biotic cycling of phosphorus in the water column
are receiving increased attention in light of discoveries that external sources
of nutrients are sometimes insufficient to explain observed rates of biological
production in lakes. The chief difficulties slowing a rapid and complete
analysis of the system are primarily technical ones that arise because fliixes
associated with production and with losses are intimately connected.

Only the outline of a picture characterizing focal processes in phosphorus
cycling is known at present with great assurance, but that outline provides a
guide for future work. A detailed kinetic description of nutrient uptake by

58

assemblages of natural phytoplankton at concentrations slightly above ambient
is a vital need. Additional research needs are measurements of nutrient
release J^ situ by the zooplankton, including not just juvenile and adult
crustaceans but rotifers and protozoans as well. Particularly important are
effects of age-structure and species composition on the flux rates anticipated
through seasons, and from year to year. The contributions of microzooplankton
have been poorly studied in proportion to their likely importance. Related to
this issue of nutrient remineralization, and of similar critical importance, is
an evaluation of mechanisms and determinants of selective grazing by
herbivorous zooplankton. That line of inquiry will probably prove central to a
dynamic description of events in the open water, because grazing limits the
standing crop of phytoplankton and thus their overall nutrient uptake capacity.
Selective grazing also represents the first step in a sequence by which
nutrients are remineralized back to a biologically-reactive dissolved form by
herbivores. Above all, the work must be rigorously quantitative. The issue
rests no longer on identifying these pathways , or even on asserting their
general importance, but rather on determining the magnitudes of nutrient and
biotic fluxes in nature and integrating them into a coherent, internally
consistent, and dynamic exposition of the structure and productivity of pelagic
lake ecosystems.

59

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68

CHAPTER IV - CONTINUED

PHOSPHORUS CYCLING THROUGH ZOOPLANKTON AS INFLUENCED

BY THEIR DIURNAL VERTICAL MIGR^^lTION, PREDATORY FEEDING,

AND POPULATION DYNAMICS

James A. Bowers

Great Lakes Research Division

The University of Michigan

Ann Arbor, Michigan

69

INTRODUCTION AND BACKGROUND

One truism in ecology that sets it apart from other disciplines is that
"everything affects everything else'* in the ecosystem. Such is the case for
the recycling of phosphorus in lakes and oceans. Virtually all biotic and
abiotic components in lakes and oceans are directly or indirectly involved with
the movement of this element. Zooplankton, an integral part of the pelagic
community, affect phosphorus recycling in lakes primarily as grazers upon
phytoplankton and predators upon themselves as well as by regeneration of
phosphorus through inefficient ingestion, egestion, and excretion. Three
aspects of zooplankton ecology, diurnal vertical migration, predatory feeding,
and population dynamics, influence these phosphorus regenerative processes.

DIURNAL VERTICAL MIGRATION

Animal migrations, their description, selective advantage, and
evolutionary origins, have long held the interest of ecologists. These regular
movements occur on seasonal, lunar, semilunar, and diurnal time scales as in
this case with the zooplankton. Hutchinson (1967) defined three patterns of
diurnal vertical migration. Nocturnal migration, the most common pattern,
begins near sunset when zooplankters ascend from daylight depths into shallower
strata, remain there through the night, and then descend to daylight depths
again at dawn. Rarely observed reversed migrations are the reverse of this
sequence. A twilight pattern results when, after the evening ascent,
a midnight sinking occurs followed by a predawn ascent. A short time later,
at dawn, the zooplankters descend again to daytime depths.

What is the relationship between vertical migration and phosphorus
recycling? When assessing the importance of diurnal migration, Longhurst
(1976) suggested that any mass migration involving such a significant fraction
of the plankton biomass has profound influences upon community metabolism. One
facet of this problem is clear; to date no completely comprehensive data set
over short time scales exists linking migration, feeding, and the vertical
distribution of food. Until this is accomplished the coupling of phosphorus
recycling to migration will remain obscure.

Diurnal migration will partially influence the spatial and temporal
dynamics of phosphorus recycling by zooplankton through the effects of
temperature, vertical distribution of food items, and diel feeding rhythms.
Zooplankters migrating during thermal stratification often experience a diurnal
rhythm in temperature. Higher temperatures will stimulate higher rates of
excretion, thus creating a diurnal cycle in excretory rates. This suggests
phosphorus excretion may be quite intense at night when migrators experience
temperatures 5^ to 10^ higher than in the daytime (Ketchum 1962). A few field
studies have suggested rhythmic excretion rates (Hargrave and Geen 1968, Eppley
ejt _al . 1973, Ganf and Blazka 1974), but detailed correlations between
temperature and excretion through time have yet to be performed.

Besides cyclical temperature regimes, migrant zooplankters also experience
radical shifts in their nutritional environment since algal and zooplankton
food items will change significantly over depth in quantity and species
composition (Bowers 1979). Only recently has this role of temporal and spatial
heterogeneity or patchiness been perceived as crucial to understanding
zooplankton nutrition (Mullin and Brooks 1976, Dagg 1977, Mayzaud and Poulet

71

1978). From earlier observations, excretion and egestion are highly correlated
to food levels. Thus phosphorus regeneration is at least indirectly tied to
migration and vertical prey and temperature distributions. Consequently,
migration may induce intense pulses of phosphorus excretion and egestion when
feeding rates are at a maximum during one diurnal cycle. Because herbivory
(Frost 1977) and predation (Kerfoot 1977a) may be selective, feeding habits
could be strongly influenced by migration and consequently linked to
assimilation and egestion. How this affects phosphorus regeneration is
presently not understood.

With higher temperature and food densities in surface waters, zooplankton
feeding is now thought to be partially a discrete event when feeding is
nocturnal (Gauld 1953, Nauwerck 1959, Mackas and Bohrer 1976, Bowers and
Grossnickle 1978) or bimodal at dusk and dawn (Haney 1973, Haney and Hall
1975). Perhaps these feeding cycles are endogenously controlled (Chisholm
et al. 1975, Duval and Geen 1975) and coupled to field observations of diel
periodicites in excretion. Laboratory feeding experiments might clarify this
relationship.

Although migrations have been thoroughly described in marine and
freshwater environments (Ringelberg 1964, Hutchinson 1967, Longhurst 1976),
their origin and the selective advantages that fostered them still remain
unsolved. Different hypotheses have been proposed involving light (Russell
1927, Harris 1953, Hairston 1976), population regulation (Wynne-Edwards 1962),
and genetic recombination (Davis 1961). More recent ideas have been based on
the trophic relationships between the migrant zooplankter and its prey and
predators. Since optimal times of food and temperature rarely coincide in time
and space, McLaren (1963) based his hypothesis on the fact that higher egg
production at colder depth strata was offset by delayed development. By
feeding only at night in the warm surface waters, copepods could maximize their
feeding efficiencies at higher temperatures and their growth and development at
lower temperatures resulting in an energy (McLaren 1963) or fecundal (McLaren
1974) bonus for the animal. Enright (1977) also believed motivation for
migration lies in the consequences of the daily alternation of temperature
experienced by migrating zooplankton in thermally stratified waters. His
metabolic model predicted a greater energy gain for growth and reproduction by
feeding at night in the warm surface water where algal production is highest
and by reducing energy demands in the daytime at the deeper (colder) depths.
Another hypothesis suggests that diurnal migration evolved from the avoidance
by zooplankton of daylight where they are vulnerable to visually oriented
planktivorous fish. Zooplankton ascend into the surface waters at night to
feed and yet avoid predation. Zaret and Suffern (1976) provided experimental
evidence supporting this hypothesis in Gatun Lake, Panama, and Fuller Pond,
Connecticut. A preferred prey of the planktivore in Gatun Lake always remained
at depths during its nocturnal migratory cycle where light intensities were too
low for the planktivore to efficiently feed. Laboratory experiments with a
daphnid prey and its planktivore from Fuller Pond showed a high correlation
between predation rates and light intensity. But, prey from Gatun Lake exhibit
this behavior in an isothermal habitat in direct contradiction to McLaren's
(1963) and Enright' s (1977) thermal hypotheses. Unfortunately, definitive and
testable hypotheses resolving this unusual behavior have not been forthcoming
(Miller 1979).

In summary, diurnal vertical migration places strong migrators on a diel
schedule where all aspects of their daily life, especially temperature,

72

feeding, and prey environment, are intimately linked to the movement. The P
regeneration mechanisms discussed earlier (Lehman, Chapter IV) indicate that
the complex migratory relationships are certainly important in phosphorus
cycling .

PREDATORY FEEDING BEHAVIOR

Most of what we know about phosphorus regeneration through zooplankton is
derived from experiments with herbivorous zooplankters due to their important
regulation of algal dynamics (see Lehman, Chapter IV). Far less is known about
predaceous zooplankters as phosphorus recyclers.

These invertebrate predators, mostly cyclopoid and calanoid copepods, are
now increasingly viewed as a major structuring force of limnetic communities
(Lynch 1977; Kerfoot 1977a, 1977b). Selecting for small zooplankton prey, they
tend to structure the community toward larger species. Cyclopoids are
herbivorous as nauplii and copepodite ins tars CI through CIII, then switch to
carnivory as adults (Gophen 1977). The adults prefer small prey with rotifers
and copepod nauplii the preferred groups (Confer 1971, Brandl and Fernando
1975a, Karabin 1978, Gilbert and Williamson 1978), although prey shape,
carapace strength, prey escape responses, and palatability also affect
selectivity (Li and Li 1979). Calanoid predators also prefer small fragile
prey (e.g. Kerfoot 1978), but recent evidence suggests some species prefer
larger prey within defined limits (Mullin 1979, Landry 1978).

Selective feeding behavior for small prey will influence phosphorus
regeneration. Inefficient consumption of larger prey or "sloppy ingestion"
presumably results in the release of prey viscera into the water when prey are
torn apart or only partially eaten and discarded by both cyclopoids (Brandl and
Fernando 1975a, 1975b) and calanoids (Anderson 1970, Bowers and Warren 1978,
Kerfoot 1978, Landry 1978). Loss of dissolved phosphorus from damaged
zooplankters is considerable (Eppley et £l. 1973, Mullin £t _al. 1975). Brandl
and Fernando (1975b) noted that cyclopoid assimilation efficiencies decreased
dramatically when prey had to be torn apart at the mouth. Thus significant
amounts of phosphorus may be lost not only during the act of feeding, but
during egestion as well. Finally, size selection for smaller prey results in
intense predation pressure on juvenile crustaceans and rotifers whose
phosphorus excretory rates are much higher than larger prey.

Rotifers are potentially an important group in nutrient recycling.
Rotifer excretion rates of dissolved phosphorus are generally higher than
crustacean species (Hargrave and Geen 1968), thus, this group is considered an
important component in phosphorus cycling due to their high metabolic rates
(e.g. Makarewicz and Likens 1979). Since rotifers are small and relatively
slow- swimming fragile prey compared to juvenile crustaceans, they are a
preferred prey of carnivorous copepods. We need more information on the food
web pathways they pass through and their quantitative role in phosphorus
excretion.

Predaceous zooplankters excrete phosphate per unit of body weight at
higher rates than herbivores (Beers 1966, Mullin £t al. 1975). This is sur-
prising since prey consumed by carnivores have higher bodily N:P ratios than
phytoplankton (Mullin et al. 1975). Perhaps they excrete more nitrogen as
organics compared to herbivores or assimilate phosphorus more efficiently than
nitrogen. For this reason excretion by carnivorous zooplankton warrants further
study.

73

POPULATION DYNAMICS OF ZOOPLANKTON

Phosphorus regeneration through excretion, egestion, and inefficient
ingestion is probably species, age, and sex specific which involves the
demography of zooplankton. Allan (1976) separated the zooplankton into two
groups: rotifers and cladocerans, and the copepods. The Rotifera and
Cladocera are "r-selectors" (MacArthur 1972), species with high intrinsic rates
of increase and adapted to seasonal or unpredictable environments. Their life
history patterns are characterized by parthenogentic reproduction and short
multivoltine life cycles. They form large highly variable populations
reacting, without time lags, to habitat changes. Copepods are "k-selectors ,"
which reproduce sexually at lower rates and have longer multivoltine or
univoltine life cycles. They respond with relatively longer time lags to
environmental changes .

Differences in population characteristics most likely translate into
significant differences in phosphorus regeneration during seasonal succession
in temperate lakes and marine waters. Rotifera and Cladocera are summer and
fall plankton groups, in which a few species will rapidly increase in numbers
for a few weeks and then decrease dramatically, to be followed by another
species group re-enacting this boom-and-bust cycle. During the rapid growth
phase of these cycles, which overlap throughout the summer, turnover rates
(production/biomass ratios/ time) are high, accounting for the majority of the
phosphorus flux in the community (e.g. Makarewicz and Likens 1979). In marked
contrast, the copepods have phosphorus flux rates several times lower. The
effects of these spurts of phosphorus regeneration on phytoplankton growth is
poorly understood. Predatory pressure on rotifers and cladocerans may
indirectly control these population and nutrient regeneration pulses by
triggering the succession of numerical decreases throughout the summer
(Threlkeld 1979). Copepods may play a dominant role in phosphorus regeneration
during the early spring when the zooplankton assemblage is dominated by
naupliar and copepodite instars which have relatively higher rates of phosphate
excretion due to their small size and rapid growth.

Another difference between these groups concerns phosphorus release rates
from fecal material, which further emphasizes the higher phosphorus flux
through rotifers and cladocerans. While copepods produce a coherent pellet
wrapped in a periotropic membrane, cladocerans and rotifers produce a pellet
which breaks up immediately upon excretion into the water. Intuitively,
nutrient release from the rapidly broken up fecal material of cladocerans and
rotifers would appear to occur at higher rates than from copepod pellets, which
gradually disintegrate from bacterial attack. The copepod pellets may serve as
an effective mechanism of P removal from the photic zone. Quantitative studies
comparing these differences are needed.

SUMMARY

Although no complete theory concerning the selective advantages of
vertical migration is imminent, feeding strategies and predator avoidance by
zooplankton will be fruitful directions for future research. Comprehensive
field efforts would further our insights into this behavior, and begin to
reveal the effects of the migrations on phosphorus regeneration. Excretion and
egestion rates have been observed to increase with higher temperatures and food

74

densities when zooplankton ascend at night to feed in the wanner and more
productive surface waters. These discrete periods of intense feeding could
induce pulses of phosphorus release. Laboratory experiments correlating
phosphorus regeneration to temperature and feeding cycles, and field efforts
coupling migration to zooplankton ingestion rates and phytoplankton patchiness,
should be a principle research objective.

Invertebrate predators play an important known role in structuring
limnetic communities, but their effect on phosphorus recycling is virtually
unknown. Two studies indicate that predatory copepods excrete phosphorus at
higher rates than herbivorous species. Inefficient ingestion of rotifers and
crustacean nauplii by predaceous cyclopoids is a potentially important but
unexplored mechanism for phosphorus release. Laboratory experiments designed
to address this behavior and predator excretion rates are needed.

Successive blooms of rotifers and cladocerans during their summer species
succession could account for a significant fraction of the phosphorus
regeneration in temperate lakes. Excretion and egestion rates during these
blooms should be highly correlated to the reproductive dynamics of these
groups. The effect of rotifers, in particular, are unclear in this succession
and, due to their very high production rates, deserve special attention.

75

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Labidocera tridpinosa . Limnol . Qceanogr . 23: 1103-1113.

Li, J. L., and H. W. Li. 1979. Species specific factors affecting

predator-prey interactions of the copepod Acanthocyclops vernalis with its
natural prey. Limnol . Qceanogr . 24: 613-626.

Longhurst, A. R. 1976. Vertical migration, pp. 116-140. In : D. H. Gushing
and J. J. Walsh (eds.). The ecology of the seas. W. B. Saunders Co.

Lynch, M. 1977. Zooplankton competition and plankton community structure.
Limnol . Qceanogr . 22: 775-777.

MacArthur, R. H. 1972. Geographical Ecology . Harper and Row.

Mackas , D. , and R. Bohrer. 1976. Fluorescence analysis of zooplankton gut
contents and an investigation of diel feeding patterns. J_. Exp . Mar .
Biol . Ecol . 25: 77-85.

Makarewicz, J. G., and G. E. Likens. 1979. Structure and functions of the
zooplankton communities of Minor Lake, New Hampshire. Ecol . Monogr .
49: 109-127.

Mayzaud, P., and S. A. Poulet. 1978. The importance of the time factor in
the response of zooplankton to varying concentrations of naturally
occurring particulate matter. Limnol . Qceanogr . 23: 1126-1143.

McLaren, I. 1963. Effects of temperature on the growth of zooplankton, and
the adaptive value of vertical migration. _J. Fish . Res . Board
Can. 20: 685-727.

. 1974. Demographic strategy of vertical migration by a marine

copepod. Am. Nat . 108: 91-102.

Miller, G. B. 1979. Comments from a nominate referee on an exchange of notes.
Limnol. Qceanogr . 24: 785-787.

Mullin, M. M. 1979. Differential predation by the carnivorous marine copepod,
Tortanus discaudatus . Limnol . Qceanogr . 24: 774-776.

, and E. R. Brooks. 1976. Some consequences of distributional

heterogeneity of phytoplankton and zooplankton. Limnol . Qceanogr .
21: 784-796.

, M. J. Perry, E. H. Renger, and P. M. Evans. 1975. Nutrient

regeneration by oceanic zooplankton: a comparison of methods. Mar «
Sci. Gomm. 1: 1-13.

78

Nauwerck, A. 1959 • Zur bestimmung der Filterierrate limnischer
planktontiere. Arch * Hydrobiol * Supp l. 25: 83-101.

Ringelberg, J. 1964. The postively phototactic reaction of Daphnia magna

Straus: Contribution to the understanding of diurnal vertical migration.
Neth . J. Sea Res . 2: 319-406.

Russell, F. S. 1927. The vertical distribution of plankton in the sea.
Biol . Rev . 2: 213-262.

Threlkeld, S. T. 1979. The midsummer d3mamics of two Daphnia species in
Wintergreen Lake, Michigan. Ecology 60: 165-179.

Wynne-Edwards, V. C. 1962. Animal dispersal in relation to social behavior .
Oliver and Boyd.

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79

CHAPTER V

FISH DYNAMICS AND PHOSPHORUS CYCLING IN LAKES

James F. Kitchell

Laboratory of Limnology

University of V/isconsin

Madison, Wisconsin

81

INTRODUCTION

Principles of the predation process and their application to phosphorus
cycling in lentic environments are summarized in other chapters of this
document (e.g. Lehman Chapter IV, Nalepa £t al. Chapter VI, and Scavia Chapter
VIII). The special case for fishes is primarily expressed in two ways.
(1) Fishes have long life histories relative to the kinetics of P cycling;
their modulating effects are second only to the "big- slow" component of
sediment-water exchange rates and, where appropriate, to the flushing rate of a
lake (O'Neill 1976). (2) Fish populations are the object of management
manipulations. For example, natural reproduction of Great Lakes salmonids and
the sea lamprey are continually altered through fisheries management: salmonid
stocks are regularly enhanced from hatcheries as recruitment although natural
reproduction is modest or non-existent; lamprey populations are maintained at
minimal levels through continuing use of lampricides. Consequently, the
quantitative effects of predation by fishes are largely controlled through
management practices .

Recent synthesis papers addressing the role of salmonid predation in
structuring Lake Michigan's fish assemblage (Stewart et al. unpublished
manuscript) and the continuing lamprey control program (Walters e_t ail . 1980)
advocate an experimental approach for management programs. If carefully
conceived, modeled, and monitored, such whole system experiments could provide
major insight into phosphorus dynamics.

The role of fishes in phosphorus cycling is the subject of two recent
reviews (Kitchell jet al. 1979, Nakashima 1980). The role of fishes may be
generally described under four major headings: (a) excretory remineralization
of P, (b) storage and translocation of P, (c) selective predation by fishes,
and (d) predator effects on community structure. Each topic will later be
discussed in greater detail. The massive dilution and mixing rates of Great
Lakes waters reduce (a) and (b) to relatively minor status. Although massive
mortality (e.g., alewife die-off s) may result in substantial fertilization, the
effects are local and of lesser significance than continuously acting and
potentially dynamic processes such as those associated with the role that
fishes play as predators.

Selective predation processes substantially determine the size-particle
distribution of prey trophic levels. Phosphorus flux rates are typically
inversely related to mean particle size. "Keystone predator" (Paine 1966) or
polyphagous predator (Hassell 1978) effects on communities become apparent as
dominant competitors are selectively removed by predation, thereby allowing
coexistence of a more diverse species assemblage including a variety of life
history strategies.

The results of experimental studies (Hurlbert jet^ _al • 1972, Cooper 1973,
Bartell 1978) have shown that consumer effects are significantly non-linear.
For example, at low predator densities, predation plays a minor role in
nutrient cycling or primary production processes. At intermediate predator
densities, however, the predation process can dramatically stimulate nutrient
cycling rates and enhance productivity per unit of nutrient. In general terms,
the mechanism is manifest as a series of predator-induced mortalities and
selection processes that reduce competitive interactions between the prey,
therebjr maintaining prey populations in the logarithmic growth phase. As a
consequence, turnover rates remain rapid, a diversity of life history
strategies is sustained, and P cycling efficiencies are high. But at very

83

intense levels of predation, prey biomass can be substantially depressed, and
local extinctions will result in reduced diversity and lower ecological
efficiencies. Production processes can be either highly variable or severely
depressed (Kitchell et al. 1979).

PHOSPHORUS REMINERALIZATION BY FISHES
Excretion

Estimates of rates of phosphorus excretion as well as attempts to measure
forms of P excreted are rare. Equally important, variability is substantial.
Furthermore, excretion and egestion are often not separated, and the several
techniques employed yield rather different results. Short-term measures
(1-3 hr) of excretion by recently captured fish generally yield high rates but
are suspect due to the unknown effects of capture stress and handling
(Nakashima 1980). Whitledge and Packard (1971) reported excretion rates of
phosphorus from anchovies in Pacific upwelling regions of 2.1 mg phosphorus g"-^
day""-^ on a dry weight basis . For a variety of marine fishes at high
temperatures (>15^C), Pomeroy and Kuenzler (1969) found an excretion range of
0.008 to 0.4 mg phosphorus g"-*- day"-^. Excretion rates similarly derived for
carp range from 0.25 to 0.27 mg phosphorus g"^ day""^ (Lamarra 1975). Based on
measures of P release during starvation (1 week) of bluegill sunfish, Koonce
and Seadler (unpublished manuscript) obtained a mean excretion rate of 0.033 mg
phosphorus g"^ day"-^ at 17^0.

Based on laboratory studies of bluegill sunfish, excretion was found to
account for 48% of the total phosphorus released by both excretion and egestion
(Koonce and Seadler, unpublished manuscript). Estimated by difference, P
excretion by yellow perch was 2-71% of total for excretion plus egestion: the
average of 15 trials was 50% (Nakashima 1980).

In general, rates of P excretion are modest compared to excretion by
zooplankton. However, a dense school of fish may create local enrichment
(Whitledge and Packard 1971). The relative importance of this special case
remains undetermined for freshwater systems.

Egestion

Due to their rapid sinking rates, fecal pellets represent a potentially
important mechanism where material consumed in the pelagic zone is lost to the
sediments. Zooplankton fecal pellets are known to be a significant vector in
marine systems (Turner and Ferrante 1979) and in some lakes (Kitchell et al.
1979). As specific consumption rates of fishes are substantially less than
those of smaller animals, the quantitative role of fecal pellet losses are
undoubtedly less, but have been estimated by Nakashima (1980) to be 10-25% of
total daily sedimentation losses in Lake Memphremagog. Herbivorous fishes in
shallow water habitats increase net sedimentation of detrital phosphorus
(Terrell and Terrell 1975) thereby increasing P concentration in surface
sediments (Stanley 1974). The relative importance of the mechanism remains
rarely estimated and will vary with the food habits of fishes and the ratio of
fish biomass to lake voltime.

The dominant form of phosphorus released through egestion and excretion

84

appears to be soluble reactive phosphorus (SRP). In a study of carp excretion,
Lamarra (1975) found 50% of total phosphorus released to be SRP which would be
available for algal uptake. Koonce and Seadler (unpublished manuscript) also
reported that SRP was a significant fraction of the total phosphorus released.
They found that the SRP fraction was consistently near 40% of the total
phosphorus released. Detailed analysis of the molecular forms of the soluble
organic fraction revealed a dominant high molecular weight fraction (>10,000
MW) and a low molecular weight fraction (^6,000 MW) .

Temperature dependence of phosphorus release by fish has not been well
studied. Koonce and Seadler (unpublished manuscript) observed that phosphorus
release declined with decreasing temperature (25^C vs. 17^C) , but that this
relationship seemed to reflect the temperature dependence of consumption. In
general, they found that the specific phosphorus release rate was a linear
function of specific consumption rate. Efficiency of phosphorus assimilation
may also vary with consumption rate. At low consumption rates, Koonce and
Seadler found an assimilation efficiency of 33%, but the efficiency approached
80% at higher feeding rates. Nakashima (1980) found that at dietary P levels
from 1.1-3.1 g P/100 g dry food, yellow perch maintained at similar rations
(1.5% of body weight day"^) and constant temperatures (12^-21^C) had relatively
similar absorption efficiencies of 71.6 + 7.2%.

In overview, carefully controlled and complete experimental determination
of P excretion and egestion rates by fish remains uncompleted. Although of
physiological interest, the rates reported to date are generally so low as to
be of little ecological consequence (Kitchell et al. 1979, Nakashima 1980)
except under the rare conditions of a large biomass of fish confined to a
relatively small volume of water (Stanley 1974, Lamarra 1975).

STORAGE AND TRANSLOCATION OF PHOSPHORUS

The P content of fish biomass is generally 3-4% of dry weight; i.e.
3-4 times that of typical invertebrate prey. Because fishes are long-lived and
have low rates of P turnover per unit mass, fishes can operate as a phosphorus
sink. In shallow productive lakes, fish biomass can be the single largest pool
of phosphorus in the limnetic zone. Kitchell £t al . (1975) estimated Lake
Wingra's fishes to contain 70% of the total pelagic P during summer months.
Nakashima (1980) estimated the Lake Memphremagog fish populations to contain an
amount of P approximately equal to that of the total sestonic P; i.e. 50% of
total pelagic P. Lake morphology and productivity are obviously important
determinants of the relative amount of P stored in fish biomass.

Migration, aggregation, and pulsed mortality of fishes followed by rapid
decomposition can yield substantial local fertilization. Approximately 50% of
the P stored in fish biomass, however, is apatitic and thereby refractory to
decomposition (Smith e£ al^. 1977). The most dramatic examples occur with
anadromous fishes .

In summary, the effects of P storage and translocation by fish appear to
be local and short-term. However, significant whole-system implications
related to enhanced recruitment can result in greater densities of
size-selective predators whose effects, as developed in the following section,
can profoundly influence P cycling in subsequent years.

85

THE EFFECTS OF SIZE-SELECTIVE PREDATION

The greatest potential importance of fishes in phosphorus cycling is
through their role as selective predators. There is substantial and diverse
evidence of the significant effect of selective predation on planktonic systems
(Hall e£ al. 1976, Porter 1977). Peters (1975) reviewed the importance of
zooplankton body size in P cycling. The dramatic effects of predation on
zooplankton assemblages have been demonstrated in systems ranging from small
ponds to the largest of lakes (e.g. Wells 1970, McNaught and Scavia 1976).

Bartell and Kitchell (1978) estimated that P flux through an equivalent
biomass of Lake Wingra zooplankton may have been increased by as much as 40
fold due to intensive size-selective predation. Enhanced productivity in
European fish ponds is ascribed to the increased nutrient cycling rates
effected through size-selective predation by cyprinid zooplanktivores (Wolny
and Grygierek 1972). Shapiro e_t al . (1975) argued that predation by
size-selective perch also enhances algal productivity but does so by
effectively releasing phytoplankton from the grazing pressure of a large and
abundant zooplankton assemblage. Thus, in some lakes, intense predation might
reduce the rate of phosphorus recycling through reduction of zooplankton
biomass.

Bartell (1978) also used Peters' (1975) model to estimate the impact on
phosphorus excretion rates of selective predation by fishes in several lakes.
Based on the data of Wells (1970), alewife reduced Lake Michigan zooplankton
size and biomass to a level such that total phosphorus flux through zooplankton
was reduced by about 50% during high (1966) versus low (1954) abundances of
alewife. Less dramatic changes were estimated for other lakes (Hutchinson
1971); six of ten showed some potential increase in P flux due to reduced mean
size of zooplankton which was not offset by equivalent reductions of biomass.

Bartell and Breck (1979) estimated from theoretical and experimental
evidence that even whole-system P cycling rates might vary up to two fold due
to the effects of selective predation. Although Nakashima (1980) discounted
selective planktivory as important in Lake Memphremagog , his study did not
consider planktivory by juvenile fishes nor the characteristics of P cycling in
a system not continuously manipulated by fish predation.

Experimental studies conducted in microcosms provide further support of
the important role of predation by fishes in phosphorus cycling. Hurlbert
e_t al . (1972) chronicled dramatic changes in the nutrient status of systems
manipulated by the addition of fish. Cooper (1973) demonstrated enhanced
primary productivity at intermediate levels of grazing by a herbivorous fish.
In general, any manipulation of zooplanktivorous fish populations will result
in altered phosphorus cycling. The direction and magnitude of change will vary
depending on the system, but will be most manifest in the epilimnion. The
effect, however, can also extend to other subsystems (e.g., hypolimnion,
littoral) due to the mobility of the consumers and the dynamics of water masses
(Kitchell £t al. 1979).

The critical unknowns revolve around a mechanistic description of
selective predation processes appropriate to a specific trophic exchange
(Vanderploeg and Scavia 1979) and some reasonable estimator(s) of potential
biomass dynamics. Optical foraging theory provides a framework for the former
(Werner 1977); the biota specific to a particular lake dimension the response
surface.

The documented selectivity of piscivores , including both other fishes and

86

man, can effectively structure fish populations and thereby alter abundances of
the zooplanktivorous forms. The same need for mechanistic, predictive
principles applies .

PREDATION, COMMUNITY STRUCTURE, AND NUTRIENT CYCLING

There exists a vast and growing literature, both theoretical and experi-
mental, addressing the role of predation in resource utilization patterns and
species packing in ecological communities. The mechanism(s) , originally
described by Paine (1966) as a "keystone" predator effect, have subsequently
been elaborated, but the central theme remains — predation can enhance prey
diversity by selective reduction of superior competitors (e.g. recent review by
Glasser 1979). Consequently, more prey species may coexist, allowing a greater
diversity of life history stages and strategies to persist within a given
habitat. The latter consequence has potentially important implications in
epilmnetic environments where resource availability may stochastically vary.
Inputs include allochthonous P loading from storm runoff or entrainment of
hypolimnetic water masses. Continuous (but intermediate) levels of predation
allow many species to coexist at population densities approximately optimal for
sustained resource utilization with modest competition, yet at population
densities sufficiently large to respond with minimal lag to pulsed increases in
resource availability. Huston (1979) discussed the role of frequency of
disturbance, competition, and predation in maintaining high diversity and
highly productive populations under dynamic equilibrium conditions. This
complexity of interacting mechanisms seems applicable at other levels of
aquatic trophic exchange (e.g. McGowan and Walker 1979), as well as in terres-
trial systems (McNaughton 1979) and aquatic systems (Cooper 1973, Porter 1977).

Huston (1979) argued that what had appeared as conflicting evidence (i.e.
some studies demonstrating increased diversity as predation pressure increased,
other studies demonstrating the reverse) simply represented either the
ascending or descending limb of a total response that is unimodal. Diversity
is enhanced at intermediate levels of predation, as represented in the sketch
below:

Diversity

Productivity per'
nutrient unit

Low

Intermediate

High

Predation pressure

87

The sketch also demonstrates a general relationship between predation
intensity and productivity per unit of nutrient similar to that derived from
experimental studies of algal production under varying intensities of grazing
pressure by a herbivorous fish (Cooper 1973). The general response is the
basic tenant of maximum sustained harvesting strategies developed for fisheries
(Larkin 1978). However, the maximum yield is rarely sustained over long time
periods due primarily to instabilities associated with stochastic effects on
recruitment and the fact that fisheries like long-lived predators have
substantial predatory inertia (Stewart et al. unpublished manuscript) relative
to the dynamics of their prey.

The greatest positive feedback of nutrient cycling rates enhanced by
selective predation occurs at intermediate levels of predation where highest
diversity is maintained. The diverse forage base supports an equally diverse
assemblage of foraging strategies. Populations are continually reduced by
predation but to levels characterized by the log phase of rate of increase and
rapid rates of turnover. Densities of various potential competitors are high
enough to better withstand reductions through stochastic effects (Chesson
1978). Mean sizes of individuals are sufficiently large to maintain high
reproductive potential yet generally much smaller than that for a dominant,
senescent population. Accordingly, size-related nutrient feedback mechanisms
operate at relatively high rates .

Lower predation rates result in lower diversity due to increased com-
petitive exclusion. Dominant competitors develop large and more senescent
populations; fewer life history strategies exist and fewer trophic linkages
yield fewer alternative nutrient feedback pathways. Consequently, stochastic
input may have a lower probability of enhancing some component of the community
or, if a negative event, a higher probability of further reducing diversity.

Very intense levels of predation also result in lower productivity due to
overexploitation. Primary prey species populations are depressed, and lags in
the growth response result in less efficient utilization of pulsed resource
availability. A higher probability of local extinctions due to stochastic
events exists due to reduced specific population sizes (Chesson 1978).

The proposed relationship between productivity per unit nutrient and
predation intensity varies with the annual cycle. Seasonal succession of
planktonic forms results in rapidly changing optima confounded by the
co-evolutionary history of organisms involved (Porter 1977, Kitchell et al.
1979). Moderate predation pressure should, however, moderate the succession
rates through reduced competitive displacement. Integrated over an annual
cycle, the cumulative effect should yield a greater total productivity and the
highest efficiency of nutrient cycling. This general mechanism is offered as a
hypothesis to be evaluated in modeling exercises and experimentation.

88

REIFERENCES

Bartell, S. M. 1978. Size selective planktivory and phosphorus cycling

in pelagic systems. Ph.D. Thesis (Oceanography and Limnology), University
of Wisconsin-Madison.

_____ , and J. E. Breck. 1979. Simulated impact of macrophyte

harvesting on pelagic phosphorus cycling in Lake Wingra, pp. 229-250. In ;
J. E. Breck e^ al. (eds.) Aquatic Plants, Lake Management and Ecosystem
Consequences of Lake Harvesting. Inst. Env. Studies, Univ. Wise. -Madison.

, and J. F. Kitchell. 1978. Seasonal impact of planktivory on

phosphorus release by Lake Wingra zooplankton. Verh . Internat . Verein .
Limnol . 20:466-474.

Chesson, P. 1978. Predator-prey theory and variability. Ann . Rev .
Ecol. Syst . 9:323-347.

Cooper, D. C. 1973. Enhancement of net primary productivity by herbivore
grazing in aquatic laboratory microcosms . Limnol . Oceanogr . 18:31-37.

Glasser, J. W. 1979. The role of predation in shaping and maintaining the
structure of communities. Am. Nat . 113:31-641.

Hall, D. J., S. T. Threlkeld, C. W. Burns, and P. H. Crowley. 1976.

The size-efficiency hypothesis and the size structure of zooplankton
communities. Ann . Rev . Ecol . Syst . 7:177-203.

Hassell, M. P. 1978. The Dynamics of Arthropod Predator-Prey Systems.
Monogr. Pop. Biol. 13, Princeton Univ. Press.

Hurlbert, S. H., J. Zedler, and D. Fairbanks. 1972. Ecosystem alteration
by mosquito fish ( Gambusia affinis) predation. Science 175:639-641.

Huston, M. 1979. A general hypothesis of species diversity. Am. Nat .
113:81-101.

Hutchinson, B. P. 1971. The effect of fish predation on the zooplankton of
ten Adirondack lakes, with particular reference to the alewife, Alosa
pseudoharengus . Trans . Am . Fish . Soc . 100:325-335.

Kitchell, J. F., J. F. Koonce, and P. S. Tennis. 1975. Phosphorus flux
through fishes. Verh . Internat. Verei n. Limnol. 19:2478-2484.

, R. V. O'Neill, D. Webb, G. Gallepp, S. M. Bartell,

J. F. Koonce, and B. S. Ausmus . 1979. Consumer regulation of nutrient
cycling. BioScience 29:28-34.

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Koonce, J. F., and A. W. Seadler. Characterization of phosphorus reminerali-
zation by the bluegill sunfish ( Leopomis macrochirus ) ; analysis of
soluble phosphorus compounds in excreta and egesta. Unpublished
manuscript, Case Western Reserve Univ., Dept . of Biology.

Lamarra, V. A. 1975. Digestive activities of carp as a major contributor to
the nutrient loading of lakes. Verh . Internat . Verein . Limnol .
19:2461-2468.

Larkin, P. A. 1978. Fisheries management - an essay for ecologists.
Ann . Rev . Ecol . Syst . 9:57-74.

McGowan, J. A., and P. W. Walker. 1979. Structure in the copepod community
of the North Pacific central gyre. Ecol . Monogr . 49:195-236.

McNaught, D. C, and D. Scavia. 1976. Application of a model of zooplankton
composition to problems of fish introductions to the Great Lakes,
pp. 281-304. In: R. P. Canale (ed.). Modeling Biochemical Processes in
Aquatic Ecosystems. Ann Arbor Science, Ann Arbor, Michigan.

McNaughton, S. J. 1979. Grazing as an optimization process: grass-ungulate
relationships in the Serengeti. Am . Nat . 113:691-703.

Nakashima, B. S. 1980. The contribution of fishes to phosphorus cycling in
lakes. Ph.D. Thesis (Biology), McGill University, Montreal, PQ, Canada.

O'Neill, R. V. 1976. Ecosystem persistence and heterotrophic regulation.
Ecology 57:1244-1253.

Paine, R. T. 1966. Food web complexity and species diversity. Am . Nat .
100:65-75.

Peters, R. H. 1975. Phosphorus regeneration by natural populations of
limnetic zooplankton. Verh . Internat . Verein . Limnol . 19:273-279.

Pomeroy, L. R. , and E. J. Kuenzler. 1969. Phosphorus turnover by coral reef

animals, pp. 474-482. In ; D. J. Nelson and F. C. Evans (eds.). Symposium
on Radioecology. USAEC (CONF-670503) .

Porter, K. G. 1977. The plant-animal interface in freshwater ecosystems.
Amer . Sci. 65:159-170.

Shapiro, J., V. Lamarra, and M. Lynch. 1975. Biomanipulation — An ecosystem
approach to lake restoration. Contrib. No. 143, Limnological Research
Center, Univ. Minnesota. Manuscript.

Smith, E. A. M. , C. I. Mayfield, and P. T. S. Wang. 1977. Colonization and
decomposition of fish bone material in natural and synthetic aqueous
solution. J. Fish. Res. Board Can. 34:2176-2184.

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Stanley, J. G. 1974. Nitrogen and phosphorus balance of grass carp,

Ctenopharyngodon idella , fed elodea, Egeria densa * Trans * Am * Fish * Soc «
103:587-592.

Stewart, D. J., J. F. Kitchell, and L. C. Crowder. Forage fishes and their
salmonid predators in Lake Michigan - past, present, and possibilities.
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Terrell, J. W., and T. T. Terrell. 1975. Macrophyte control and food habits
of the grass carp in Georgia ponds. Verb . Internat . Verein . Limnol .
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Turner, J. T., and J. G. Ferrante. 1979. Zooplankton fecal pellets in
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Vanderploeg, H. A., and D. Scavia. 1979. Two electricity indices for feed-
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Walters, C. J., G. Spangler, W. J. Christie, P. J. Marion, and J. F. Kitchell,
1980. A synthesis of knowns, unknowns, and policy recommendations from
the Sea Lamprey International Symposium. Can . _J. Fish . Aquat . Sci .
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Wells, L. 1970. Effects of alewife predation on zooplankton populations in
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Whitledge, T. E. , and T. T. Packard. 1971. Nutrient excretion by anchovies
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Wolny, P., and E. Grygierek. 1972. Intensification of fish pond production,
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91

CHAPTER VI

THE BIOLOGICAL COMPONENT OF PHOSPHORUS EXCHANGE
AND CYCLING IN LAKE SEDIMENTS

Thomas F. Nalepa

Great Lakes Environmental Research Laboratory
NOAA, Ann Arbor, Michigan

and

David S. White and Catherine M. Pringle

Great Lakes Research Division and

School of Natural Resources

University of Michigan

and

Michael A. Qviigley

Great Lakes Environmental Research Laboratory
NOAA, Ann Arbor, Michigan

93

INTRODUCTION

The exchange and cycling of phosphorus between lake sediments and the
overlying waters is a complex process that occurs in several stages: input via
particulate deposition, biotic mineralization and diagenetic chemical
transformation, release into the interstitial waters and, finally, release
through the sediment -water interface. The rate at which each of these stages
proceeds is dependent upon the interaction of numerous physical, chemical, and
biological variables. This review will summarize what is known of the
biological aspects of the regenerative process.

Most of what is known of the relationships between biotic activities and
sediment P flux has been gained only in the past decade. The classical concept
of sediment P exchange viewed the sediments as a sink; release of P into the
overlying waters only occurred during periods of low oxygen concentrations, as
in the hypolimnion of eutrophic lakes during summer stagnation, and was
intimately related to the chemical kinetics of iron phosphate compounds
(Mortimer 1941, 1942; Hayes e£ al . 1952; Stumm and Leckie 1971). Recent
evidence, however, indicates that P release also occurs when dissolved oxygen
concentrations are high. Although release under aerobic conditions is about
ten times less than under anaerobic conditions (Burns and Ross 1972, Holdren
and Armstrong 1980), the significance of aerobic P release should not be
minimized. In the Great Lakes, Mortimer (1971) predicted that significant P
release would only occur when oxygen concentrations were below 2 mg/L, and
Burns and Ross (1972) found anoxic regeneration of P in central Lake Erie only
when oxygen concentrations were below .6 mg/L. Such low oxygen levels are
found only in a limited area of the Great Lakes. Aerobic conditions exist
throughout the nearshore area and sediment P release here would be
well-circulated and available for algal uptake. Under some conditions, as
shown by experiments on deepwater Lake Ontario sediments, aerobic release may
even equal anaerobic release (Bannerman et al. 1975).

The exact mechanisms of aerobic P release are virtually unknown, but
appear related to the activities of benthic organisms. Wood (1975) suggested
that, of several mechanisms responsible for materials movement across the
sediment-water interface, ordinary diffusion was the least important while
biotic activities were of considerable importance. Holdren and Armstrong
(1980) found that, of the factors they investigated, the presence of benthic
organisms had the greatest effect on sediment P release rates. Lee et al .
(1977) attributed aerobic P release to the microbial mineralization of P
compounds .

Traditionally, benthic organisms can be divided into several categories
based on relative size: the macro-, meio-, and microbenthos. In the Great
Lakes, typical representatives of each of these categories are oligochaetes ,
Pontoporeia , chironomids, and pelecypods (macrobenthos); nematodes,
cladocerans, and copepods (meiobenthos) ; bacteria and algae (microbenthos).
The biotic activities of each of these benthic groups affect the uptake and
release of sediment P, and although each activity mediates sediment P in a
unique way, each is intimately related to the activities of the other groups.
The biotic role can be broadly defined by the following activities:
bioturbation, uptake and excretion, and organic mineralization.

95

MACROBENTHOS AND BIOTURBATION

Bioturbation is the mechanical mixing of the sediments by the burrowing
and feeding activities of the benthos. Though meio- and microbenthos may have
some role in bioturbation, most of the actual mixing process is a function of
the macrobenthos, particularly those taxa with body dimensions greatly
exceeding the sizes of the sediment particles. Such stirring activities lead
to both physical and chemical changes that alter exchange processes.
Bioturbation increases mixing and exchange of interstial water, facilitates the
import of oxygen, affects Eh and pH profiles, and alters sediment structure.
These changes, in turn, affect the rate of phosphorus release either by
increasing exchange along concentration gradients or by changing the solubility
of phosphorus compounds.

Modes of bioturbation are species specific. The impact of a particular
species depends on its size, feeding system, and depth of penetration. In the
nearshore Great Lakes, the most abundant macrobenthic forms are the oligochaete
worms, especially in enriched areas. These deposit-feeders typically ingest
sediments below the surface (2-5 cm), then defecate at the sediment-water
interface. Thus, large volumes are moved upward, effectively "recycling"
buried deposits for further aerobic mineralization and nutrient release and
absorption of new materials. The fecal material differs in its chemistry and
structure from the underlying sediments in ways that enhance microbial growth
and mineralization (Brinkhurst 1972). Mixing of the upper sediments is
complete, as evidenced by experiments using radioactive tracers (Robbins et al.
1977) and pollen grains (Davis 1974a). The amount of sediment reworked can be
put in perspective by comparing it to rates of deposition. In the profundal of
Lake Huron, the annual amount of sediment reworked by the lumbriculid oligo-
chaete Stylodrilus heringianus in the upper 3 cm of sediment was four times
greater than the annual sedimentation rate (Krezoski et al. 1978), and in
western Lake Erie the amount reworked by tubificid oligochaetes was ten times
greater (Fisher 1979). Under optimum conditions and at densities found in some
areas of the Great Lakes, laboratory experiments have shown that tubificids are
capable of reworking the top 5 cm of sediment every 2 weeks (McCall and Fisher
1977).

The physical effects of sediment reworking by oligochaetes can be
profound. McCall and Fisher (1980) found that virtually all sediment
properties that they examined - sediment grain size, settling velocity,
erodibility, porosity, and permeability - were altered by factors of at least
two to sometimes ten or more by the presence of tubificids. The continuous
deposition of fecal material at the sediment surface produces an uncompacted
layer that is easily resuspended by bottom currents. The implication of these
activities would be greatest in the nearshore where the extent of
wind-generated sediment transport and the physical regeneration of P have been
aptly demonstrated by Lam and Jaquet (1976).

Besides altering the character of the sediments by their burrowing and
feeding, tubificids actively pump water through their burrows, as related to
respiratory requirements. Although such activity does not represent bioturba-
tion in the strict sense, these movements increase exchange between nutrient-
rich interstitial waters and the overlying waters, leading to increased
diffusion along concentration gradients. In polluted Toronto Harbour, which
has an extremely high tubificid density, the amount of water circulated through
the sediments was estimated at between 184 and 291 mL m-^ hr--^ (Wood 1975).

96

In the offshore waters of the Great Lakes, the dominant organism is
Pontoporeia hoyi. This amphipod randomly burrows in the top 1-2 cm of sediment
and makes brief excursions into the overlying waters. These activities
resuspend sediment particles and prompted Robbins e£ _al. (1979) to model them
as an "eddy diffusion" process. In the Baltic, a highly turbid zone extending
above the bottom has been observed in areas with high densities of Pontoporeia
af finis (Ankar 1977). The effect of this sediment mixing on phosphorus flux
can be derived from laboratory experiments; sediment agitation has consistently
led to increased phosphorus release.

Other macrobenthic components include chironomids and pelecypods. Many
species of the former group build vertical burrows extending from the sediment
surface to a depth of 2-3 cm. While these larvae do not continually move
particles as do the tubificids, they pump more water through the sediments
since circulation is required for both feeding and respiration. Tessenow
(1964) reported that chironomids created oxidized mud tubes to a depth of 10 cm
(most species probably affect only the top few cm) and that this activity
greatly increased the release of dissolved silica. In addition, the emergence
of adult chironomids has an immediate and pronounced effect on nutrient release
rates when densities are high (Holdren and Armstrong 1980). Sphaeriids
(fingernail clams) are found throughout the Great Lakes, sometimes in great
abundance. Although surface dwellers, they are capable of extending their foot
2.5 cm into the sediments. McCall _££ a^. (1979), in a series of laboratory
experiments, found that a large pelecypod species, commonly found in western
Lake Erie, significantly altered sediment porosity and induced incomplete
sediment mixing down to 20 cm.

MACROBENTHOS AND P EXCRETION

The extent to which the macrobenthos mediate nutrient exchange through
digestive processes is not well-defined. M though the excretion rate of
phosphorus by zooplankton is well documented (See Chapter IV), studies to
determine the excretion rate of benthic invertebrates are few and limited to
large marine species (Johannes 1964a, Kuenzler 1961). The potential impact of
organisms converting organically bound phosphorus to more soluble forms and
depositing particulate phosphorus via faecal pellets can be substantial. This
is true especially for deposit feeders, considering the high rate at which
material is passed through their gut. For instance, Brinkhurst (1972)
estimated that, of 833 tons of nitrogen entering Toronto Harbour each year, 113
tons circulated through the tubificid population alone.

In studies where sediment phosphorus release is directly attributed to
faunal activities, it has not be determined whether the release resulted from
bioturbation or excretion (Graneli 1979, Gallepp 1979, Holdren and Armstrong
1980). However, Gallepp (1979), using a regression equation relating dry
weight to P excretion developed from marine species (Johannes 1964b) , estimated
that chironomids could account for all the phosphorus released during the
course of his experiments .

97

MACROBENTHOS AND SEDIMENT P RELEASE

In experiments specifically designed to measure the impact of macrobenthic
activities on sediment P flux under aerobic conditions (paired comparisons with
and without organisms) , the organisms either decreased P in the overlying
waters (Davis et^ al* 1975), caused no change (McCall and Fisher 1980), or
increased it (Gallepp 1979, Graneli 1979, Holdren and Armstrong 1980). When
organisms did increase phosphorus release, the increase over background levels
was quite variable (Table 1). These results do not indicate inconsistencies in
macrobenthic activity, but reflect the complexity of sediment P dynamics. P is
present in lake sediments as organic, absorbed, iron-, aluminum-, and
calcium-bound phosphates. The proportion of P available for release depends on
the compounds present and on the kinetics of such chemical exchange reactions
as acid- base, precipitation, complexation, oxidation-reduction, and sorption
(Lee 1970). Since most of these reactions are complex functions of Eh and pH,
P compounds can be divided into two categories, redox-sensitive or
redox-insensitive (Davis et al. 1975). For example, while insoluble Fe (III) P
is reduced to soluble Fe (II) P under low oxygen conditions and suitable pH
(redox-sensitive), ion exchange of P between the sediments and water is not
responsive to Eh changes (redox-insensitive). Bioturbation increases sediment
Eh, which results in a net uptake by the sediments if most P is redox
sensitive, but it also increases interstitial and overlying water exchange,
which results in a net release by the sediments if most P is redox-insensitive
or a result of biotic mineralization and excretion. However, if faunal
activities lead to sediment P release, it does not necessarily follow that
redox-sensitive P is not involved. The amount of redox-sensitive P that is
fixed or released is a function of sediment composition. Since the binding
capacity of redox-sensitive compounds is decreased at high pH levels, increases
in sediment Eh would be less important in calcareous sediments. Both Graneli
(1979) and Holdren and Armstrong (1980) found that chironomid activities
increased P release to a greater extent in calcium-rich deposits.

In general, it seems that increased Eh as a result of bioturbation is not
as important in binding P as increased water exchange is in releasing it.
Williams and Mayer (1972) questioned the effectiveness of the oxidized
microzone as a barrier to P release when it is unconsolidated and has a high
water content. It is interesting to note that all studies with chironomids
have shown an increase in sediment P release, while all studies with tubificids
have shown a net uptake of P by the sediments (Davis ejt al. 1975) or no change
(McCall and Fisher 1980), even though both chironomids and tubificids increase
the import of oxygen and alter Eh and pH profiles (Edwards 1958, Schumacher
1963, Hargrave 1972, Davis 1974b, McLachlan and McLachlan 1976). This may be a
result of the fact that chironomids probably pump more water through the
sediments than do tubificids.

MEIOBENTHOS

Very little is known of the role of the meiobenthos in nutrient regenera-
tive processes. This is particularly true in the Great Lakes where only one
study (Nalepa and Quigley 1980) has quantified these forms. Because of their
relatively small size and body mass when compared to the macrofauna, the
ability of meiobenthos to cause significant changes in sediment character

98

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99

(bioturbation) has been considered slight. However, bioturbation (and
excretion) can be potentially important because of their high metabolic rates
and great densities. In the only study to document bioturbation by the
meiobenthos, nematodes and associated fauna developed an oxidized surface layer
.5-1.5 cm thick in anoxic sediments within a few hours (Cullen 1973). Unlike
the macrofauna, nematodes are able to tolerate, even thrive, under anoxic
conditions (Ott and Schiemer 1973). Considering the importance of anaerobic
chemical processes in sediment P release, the impact of nematode activities
under such conditions can be significant.

There is evidence that the meiofauna, particularly nematodes, are
important in enhancing the mineralization of organic detritus by bacteria. In
laboratory experiments, the mineralization rate of detritus in microcosms with
macrobenthic detritivores and nematodes was twice that of microcosms with
macrobenthic detritivores alone (Lee et al. 1975, Tenore et al. 1977).
Interestingly, the presence of nematodes also increased the net incorporation
of detritus by the macrobenthic detritivores.

Many species of meiofauna, e.g. copepods and cladocerans, freely enter the
sediments from the water-column, particularly in response to diurnal light
changes. At times, "planktonic" forms are more abundant in the sediments than
in the overlying waters (Nalepa and Quigley unpub. data). Wetzel (1975)
suggested that these sediment excursions are for feeding purposes. In
addition, many benthic taxa enter the water column (Wiley and Mozley 1978).
Frequent movement of organisms across the sediment-water interface may
represent an additional pathway for nutrient release — uptake in the sediments
followed by release (excretion) in the water column. The impact of these
activities would be greatest in the nearshore area where the water volume to
sediment ratio is small, and the density of organisms per unit volume is high.

MICROBENTHOS

Bacteria play a diverse role in sediment-nutrient flux. In addition to
influencing the solubility of P compounds with their metabolic end products,
bacteria form an important link between the nutrients locked in detritus and
the nutrient uptake and release by the macro- and meiobenthos. It is generally
recognized that the bacteria are responsible for most, if not all, mineralized
organic input. They are capable of assimilating a variety of organic compounds
which other heterotrophs cannot utilize (Nielson 1962). Since bacteria, in
turn, are utilized as a food source by other benthic forms, they represent
perhaps the most important link in the cycling of sediment nutrients.

The bacteria-benthic fauna association and its implications to nutrient
flux are more complex than a simple food-chain interaction. Macro- and
meiofaunal activities benefit bacteria growth and, thus, stimulate organic
decomposition by (1) improving the import of oxygen, (2) breaking down larger
particles, thereby improving the substrate for colonization, and (3) feeding on
bacteria and keeping them from reaching self-limiting numbers. Laboratory
experiments consistently indicate improved bacterial activities in the presence
of faunal grazers (Hargrave 1970, Fenchel 1972, Lopez ejt al. 1977, Lee et al.
1975, Tenore _et £l . 1977). Zvetkova (1972) noted that the rate of biochemical
oxidation of organic matter increased 1.5 to 2 times in the presence of
tubificids. Provini and Marchetti (1976) found that the total oxygen
consumption of sediments was seven times greater than the sum of the oxygen

100

prising considering the multitude of
ase, i.e. binding capacity of

uptake of sediments and tubificids measured separately. This difference was
attributed to the increased microbial activity resulting from sediment
irrigation and uplifting by the tubificids.

Tubificids (and probably other Great Lakes benthic groups as well) have a
remarkable preference for special bacteria groups and possess the ability to
detect and discriminate them among others (Brinkhurst and Chua 1969, Wavre and
Brinkhurst 1971, Chua and Brinkhurst 1973). How this selective feeding affects
the rate of organic decomposition is unkno^wn.

When sedimented organic material is mineralized, a portion is returned in
soluble form to the overlying waters, and part remains in the sediment. Burns
and Ross (1972) estimated that in Lake Erie, under aerobic conditions, 25% of
sedimented P is returned to the water coluinn and 75% remains complexed in the
sediments. The sediment P either accumulates and is held by bacteria, or is
absorbed on the sediments. Pomeroy et _al. (1965) found that both of these
pools are involved in equilibrium exchanges. In undisturbed sediments, most of
the P exchange is a physical sorption reaction; bacteria beneath the
sediment-water interface also exchange P, but the exchange is with the
interstitial waters which, in turn, only slowly diffuse upward. When the
sediments are disturbed (as during bioturbation) or scoured, bacterial P freely
exchanges with the overlying waters .

During long-term laboratory experiments, the addition of antibiotics to
inhibit bacteria has caused either a net increase (sediment desorption)
(Kamp-Neilson 1974), no change (Pomeroy ejt ail. 1965, Fillos 1977), or a net
decrease (sediment sorption) (Hayes and Phillips 1958) of P into the overlying
waters. Such variable results are not sur^
factors that can potentially affect P relec
sediments, concentration gradients, degree of sediment disturbance, etc. It is
unfortunate that none of these studies documented faunal densities.

Epipelic algal communities of the sediment-water interface may provide a
driving force for more rapid flux of nutrients (Lee 1970). The flux may
stimulate planktonic algal growth, a well documented phenomenon (Gahler 1969;
Goltenaan et al. 1969; Fitzgerald 1970a, 1970b; Porcella et al. 1976). For
laboratory algal culture studies in which calcium hydroxyapatite was used as
the phosphorus source, the algae acted as a "sink" by driving reactions in the
direction of material transfer from apatite to the organism (Gerhold and
Thompson 1969). Williams _et al . (1980) reported that P uptake by algae was
related to the amount of nonapatite inorganic P in the sediments. Ripl and
Lindmark (1979) demonstrated that P supplies within laboratory cultures of
S e lens t rum capricornatum were controlled partially by both sediments and algal
abundance. Phosphorus release from the sediment was enhanced when algae were
present, and it was concluded that mutual interactions between algae and
bacteria largely control the form of nutrient release. Porcella jet al^. (1970)
noted that the relative abilities of different sediment types to support algal
growth were related to the amount of available phosphorus in the sediments.
Further, physical and chemical fluctuations induced by algal populations
appeared to directly affect phosphorus flux from the sediments. In the
Porcella ejt al . (1970) study, the development of a thick mat of Oscillatoria on
the sediment resulted in (1) the development of anaerobic conditions, (2) the
disruption of mixing with overlying waters, (3) a higher transfer rate of
phosphorus from the sediments, and (4) the providing of a source of organic
materials (secreted by the algal cells) which served as a substrate for other
organisms. The role of algae actually occurring in the sediments (epipelic
types, primarily diatoms) in phosphorus cycling is unknown.

101

FISH

Although not part of the benthos in the strict sense, bottom-feeding fish
influence sediment P release by stirring the sediments and selectively preying
on benthic forms. Grygierek ejt _al. (1966) reported that blooms of various
phytoplankton species tend to last longer when bottom-feeding fish are present.
Hrbacek ^ _al. (1961) noted that fish removal from a lake caused a decrease in
chlorophyll concentration. Needham (1966) similarly observed that the
Chlorophyta decreased steadily after the removal of bottom-feeding fish.
Kitchell ejt _al. (1975) concluded that bottom-feeding fish excreting relatively
low but constant levels of phosphorus may act to stabilize planktonic
communities which would otherwise exhaust available phosphorus. Likewise,
Lamarra (1975) viewed the digestive activities of carp ( Cyprinus carpio) as a
major contributor to the nutrient loading of lakes. Although fish excrete
phosphorus, they probably have a greater impact on the phosphorus cycle by
structuring the benthic community through selective predation (Bartell and
Kitchell 1978).

BENTHOS-SEDIMENT INTERACTIONS IN NUTRIENT CYCLING MODELS

In spite of the evidence that benthic animal activities affect nutrient
exchange processes, few models account for the consequences of these
activities. Berner (1977) avoided modeling exchange processes in the top 10-20
cm of sediments where animal activities were most prevalent, and Vanderborght
et al . (1977a, 1977b) chose to model processes in areas having high surface
currents and correspondingly low animal densities. Such reluctance to include
biotic components in modeling efforts is due to an appreciation of the
important role of the benthos in sediment-water exchange processes and the lack
of detailed information at hand. To date, one group of researchers (Goldhaber
^ ^. 1977) included the effects of bioturbation in a steady-state diagenetic
model of sulfate reduction and diffusion. Similarly, the inclusion of animal
activities will lead to more reliable models of nutrient cycling in the Great
Lakes .

SUMMARY

The importance of benthic organisms in the cycling and exchange of
phosphorus and other nutrients between sediments and overlying water has only
been recognized in the past decade. Recent evidence indicates that aerobic
sediments are not a complete sink for phosphorus, and that more phosphorus is
regenerated than can be accounted for by diffusive processes alone. The
relationship between the biotic component and sediment phosphorus flux is
complex, partly as a result of the functional interactions between the various
benthic organisms, and partly as a result of the complexity of sediment
phosphorus kinetics.

At this point, three major functions of the benthos in phosphorus cycling
and regeneration have been defined. The first, bioturbation, or the physical
mixing of sediment by the larger organisms, leads to changes in porosity,
increases interstitial water and oxygen exchange, and alters Eh and pH
profiles. In turn, these changes may alter phosphorus cycles, depending on the

102

forms of phosphorus present and the physical and chemical character of the
sediment* Both oligochaetes and chironomids play a major role in this process.
Second, such biological activities as ingestion, assimilation, and excretion
transform phosphorus compounds; however, actual measurements of P flux as a
result of these activities are non-existent. The third function of benthos is
the mineralization of organic compounds and the subsequent release of P.

Laboratory experiments sepcifically designed to determine the impact of
organism activities on sediment P flux, e.g. comparisons with and without
organisms, have given conflicting results and are an indication of the number
of variables that can potentially alter exchange processes.

103

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Lam, D. C. L. , and J. M. Jaquet. 1976. Computations of physical transport and
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, W. C. Sonzogni, and R. D. Spear. 1977. Significance of oxic vs.

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experimental approach toward understanding the role of meiofauna in a
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McLachlan, A. J., and S. M. McLachlan. 1976. Development of the mud habitat
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macro- and meiobenthos in southeastern Lake Michigan. NOAA Data Report
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invertebrates in North Dakota lakes. Invest . Fish Control . 4: 16.

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1^7

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108

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109

CHAPTER VII

THE ROLE OF MACROPHYTES IN PHOSPHORUS CYCLING

Kwei Lin

Great Lakes Research Division

The University of Michigan

Ann Arbor, Michigan

111

INTRODUCTION

One of the characteristics of the littoral zone of lakes is the growth of
aquatic macrophytes . The littoral zone, in relation to the size of the pelagic
region, varies greatly among lakes depending on the geomorphology of the basin.
Because the littoral zone intercepts both the terrestrial and aquatic systems,
its chemical and biological features are inherently influenced by both systems.
Rich nutrients received from the drainage basin often contribute to intensive
biological activity. Littoral flora constitute the major source of synthesis
of organic materials that regulate the biological productivity and metabolism
in the littoral zone.

Aquatic macrophytes constitute the macroscopic forms of aquatic vegetation
including macroalgae, pteridophytes (mosses and ferns), and the angiosperms
(flowering plants). According to their growth forms, aquatic macrophytes can
be differentiated into submergents, emergents, and free floaters. The biology
of aquatic macrophytes has been a subject of many investigations and reviewed
in great detail by Gessner (1959), Sculthorpe (1967), and Hutchinson (1975).
One of the most significant roles of aquatic macrophytes in the aquatic
ecosystem is their involvement in the nutrient cycling in the littoral zone.

PHOSPHORUS CONTENT IN AQUATIC MACROPHYTES

The content of P in aquatic macrophytes varies widely depending on the
species, growth cycle of plants, and availability of the nutrient in the growth
medium. To evaluate the nutrient status of the aquatic vegetation, the tissue
analysis technique was introduced to establish the critical, deficient, and
luxury concentrations of nutrient (Gerloff and Krombholz 1966, Fitzgerald
1972). Tissue analysis requires establishing the critical concentration of P
or other elements of interest for each species: that concentration is the
minimum level for maximum growth. Below this level is the deficient zone in
which the plant yield is proportional to its P content. Above this level is
the zone of luxury consumption in which the P content in the organism increases
but the growth does not. The status of element content in the plant indicates
the nutrient availability in the growth media.

In experiments with several species of aquatic plants, Gerloff and
Krombholz (1966) concluded that the critical P level of 0.13% was applicable to
aquatic vegetation in Wisconsin lakes. Moreover, they reported that the
aquatic macrophytes in eutrophic and mesotrophic lakes in Wisconsin contained P
above the critical value and those in oligotrophic water below that value. The
P content, however, may change considerably during the growth season (Adams and
McCracken 1974). Reimer and Toth (1968) analyzed vegetative parts of 37
species of aquatic plants and found that the P content ranged from 0.1 to
1.95%, with a mean of 0.45%. Bernatowicz (1969), however, reported a lower
range (0.01-0.61%, with a mean of 0.17%) of P from 16 species.

UPTAKE AND RELEASE OF P

The uptake and release of P by aquatic macrophytes have been a subject of
numerous studies (e.g., Gerloff 1969, Fitzgerald 1972, McRoy and Barsdate 1970,
Denny 1972, McRoy £t al. 1972, DeMorte and Hartman 1974, Peverly and Brittain

113

1978) • It is a well-known phenomenon that the submerged macrophytes can absorb
P by the shoot-foliage portion from the water and by the root-rhizome system
from the substratum. The efficiencies of P uptake by those two portions of
plants are quite variable among species (Sculthorpe 1967).

In addition to active excretion of P by living macrophytes, leaching from
dead biomass is rapid. Even under sterile conditions a loss of from 20 to 50%
of total P may occur in a few hours. Phosphorus released from decaying
macrophytes is mineralized quickly, being either rapidly utilized by bacterial
and algal populations or incorporated into the sediments (Nichols and Keeney
1973).

The dynamics of phosphorus cycling in relation to littoral vegetation was
repeatedly demonstrated earlier using radioactive -^^p (e.g. Hutchinson and
Bowen 1947, 1950; Cof f in et al. 1949; Hayes jet al . 1952; Hayes and Phillips
1958). Their results indicated that the P added to the surface water was
rapidly absorbed by littoral vegetation and phytoplankton. Consequently the P
was slowly released to the surrounding water by the plants and rapidly released
by the epipelic algae on the plants.

PHOSPHORUS IN GREAT LAKES MACROALGAE

The distribution of aquatic macrophytes in the Great Lakes littoral zone
is relatively limited, and their role in nutrient cycling in the whole lake
system appears insignificant. Certain species of filamentous algae, however,
are common to the lower Great Lakes, and excessive growth of these algae often
occurs in nutrient-rich water. Cladophora glomerata , a highly branched green
alga, constitutes the most important aquatic macrophyte in many areas of the
Great Lakes littoral region (International Joint Commission 1975), particularly
in western Lake Erie (Taft and Kishler 1973), the southern shore of Lake
Ontario (Wezernak et al. 1974), and the southwestern portion of Lake Michigan
(Herbst 1969, Lin and Blum 1973). The most comprehensive information on
Cladophora distribution is available for Lake Ontario. In western Lake
Ontario, the Cladophora growth was estimated to occupy an area of 2,300 acres
within 3-meter contour between Toronto and Hamilton. From Rochester east to
Stony Point, Cladophora covers 79% of the bottom in a band 350-m wide. In Lake
Erie, the dense Cladophora growth occurs on the southern shore of the eastern
basin where a 35 mi-^ growth bed was estimated. The total area of Cladophora
growth on the U.S. shore of Lake Erie was estimated at 350 mi^ (International
Joint Commission 1975). The average Cladophora biomass from samples collected
in Lake Ontario was 168 g/m^ dry wt and in Lake Erie was 100 g/m^ (see Neil
1974). The extensive Cladophora occurrence in those areas may have a
significant impact of P cycling in those littoral zones. In general, the
distribution of Cladophora in the lower Great Lakes is primarily limited by
light, substrate availability, and nutrient level in the lake water.

Nutrient supplies and limiting nutrients for Cladophora growth in the
Great Lakes have been the subjects of many investigations (Neil and Owen 1964;
Herbst 1969; Fitzgerald 1970; Lin 1971,1977; Taft and Kischler 1973; and
Gerloff and Fitzgerald 1976). Extensive growth of Cladophora usually occurs in
the areas receiving nutrients from municipal waste water discharges which
contain a high concentration of P. Although the growth rate and distribution
^^ Cladophora have been shown to be closely related to phosphorus inputs, the
key phenomenon of phosphorus uptake kinetics in relation to growth rate is not

114

well known. It is well dociomented that Cladophora can take up P from an
external source and build up an internal pool as a result of luxury uptake
(Gerloff and Fitzgerald 1976, Lin 1977). Recently, Auer and Canale (1980)
observed that internal P pools may regulate the uptake of P through a
saturation feedback mechanism, i.e. the apparent maximum P uptake rate
decreases with increasing internal pool size. Under P-limiting conditions, the
P uptake rate increases with increasing external P concentration. While the P
uptake kinetics are expected to follow a hyperbolic function, the saturation
concentration for uptake is often greater than the P value observed in natural
systems. Compared to planktonic algae, Cladophora has been shown to have
certain disadvantages in P uptake. For example, its half saturation constant
is much greater than that of planktonic algae, and the maximum P uptake rate
(on dry weight biomass basis) is much smaller than that measured for several
species of phytoplankton.

The critical cellular P concentration required for maximum growth of
Cladophora is 0.06% of dry weight (Gerloff and Fitzgerald 1976). This value is
relatively lower than that of other aquatic macrophytes (Gerloff and Krombholz
1966). Gerloff and Fitzgerald (1976) also showed that samples of Cladophora
collecteid from all the Great Lakes except Lake Superior contain P con-
centrations ranging from 0.07 to 0.82%, indicating that most of the Cladophora
growth in the Great Lakes is P sufficient. Unlike rooted aquatic macrophytes
that can absorb their phosphorus from the substrate by root systems (Carignan
and Kalff 1980), Cladophora must obtain its nutrient from surrounding water
through the filaments. Since this alga noinnally occurs in a shallow littoral
region where sediment resuspension prevails, it may be able to use condensed P
originated from sediments. Lin (1977) reported that Cladophora could utilize
pyro- and tripoly-phosphate efficiently because the algae can enzymatically
hydrolyze condensed P on the cell surface and extracellularly.

In regard to the efflux of P in Cladophora , Auer and Canale (1980)
observed that algae excrete P when cells with high internal P pools are placed
in a low P environment. Furthermore, they showed that the half-saturation
constant for P excretion is approximately 20% of the maximum pool size. In
addition to P release by live cells, the nutrient regeneration through the
decay of Cladophora biomass is expected to be a significant process in which a
massive quantity of this alga periodically becomes detached from its substrate.
Unfortunately, very little investigation has been done on P release from
Cladophora and its effect on nutrient cycling in the Great Lakes littoral zones
where this alga is an important vegetation.,

Besides Cladophora , other filamentous macroalgae common to the Great Lakes
are Bangia atropurpurea and Ulothrix zonata . Although they seldom develop to
significant standing crops, the recent invasion of Bangia in the Great Lakes
(Lin and Blum 1977) may have indicated a threshold change in certain nutrient
regimes in Great Lakes water. Nutrient requirements for these algae remain
uninvestigated .

115

REFERENCES

Adams, M. S., and M. D. MeCracken. 1974 • Seasonal production of the
Myriophyllum component of the littoral of Lake Wingra, Wisconsin.
J. Ecol. 62:457-465.

Auer, M. T., and R. P. Canale. 1980. Phosphorus uptake dynamics as related

to mathematical modeling of Cladophora at a site on Lake Huron. J_. Great
Lakes Res . 6:1-7.

Bernatowicz, S. 1969. Macrophytes in the Lake Warniak and their chemical
composition. Ekol . Pol . Ser . A. 17:447-467.

Carignan, R. , and J. Kalff. 1980. Phosphorus sources for aquatic weeds:
Water or sediments? Science 207:987-988.

Coffin, C. C, F. R. Hayes, L. H. Jodrey, and S. G. Whiteway. 1949. Exchange
of materials in a lake is studied by the addition of radioactive
phosphorus. Can . £. Res. (Ser. £) 27:207-222.

DeMorte, J. A., and R. T. Hartman. 1974. Studies on absorption of -^^P,
5^Fe and ^^Ca by water milfoil ( Myriophyllum exalbescens Fernald) .
Ecology 55:188-194.

Denny, P. 1972. Sites of nutrient absorption in aquatic macrophytes.
J. Ecol. 60:819-829.

Fitzgerald, G. P. 1970. Evaluation of the availability of sources of
nitrogen and phosphorus for algae. jJ. Phycol . 6:239-247.

. 1972. Bioassay analysis of nutrient availability.

pp. 147-169. In : H. E. Allen and J. R. Kramer (eds.). Nutrients in
Natural Waters. Wiley and Sons, New York.

Gerloff , G. C. 1969, Evaluating nutrient supplies for the growth of aquatic
plants in natural waters, pp. 537-555. In : Eutrophication: Causes,
Consequences, Correctives. National Academy of Sciences. Washington, D.C.

, and G. P. Fitzgerald. 1976. The nutrition of Great Lakes

Cladophora . EPA, Duluth, Minnesota. 110 pp.
, and P. H. Krombholz. 1966. Tissue analysis as a measure of

nutrient availability for the growth of angiosperm aquatic plants.
Limnol. Oceanogr . 11:529-537.

Gessner, F. 1959. Hydrobotanik. Die Physiologischen Grundlagen der

Pflanzenverbreibing in Wasser. II. Stoffhaushalt. Berlin. VEB Deutscher
Verlage der Wissenschaf ter .

116

Hayes, F. R. , and J. E. Phillips. 1958. Lake water and sediment. IV. Radio-
phosphorus equilibrium with muds, plants, and bacteria under oxidized and
reduced conditions. Limnol . Qceanogr . 3:459-475.

, J. A. McCarter, M. L. Carmeron, and D. A. Livingstone. 1952.

On the kinetics of phosphorus exchange in lakes. £. Ecol . 40:202-216.

Herbst, R. P. 1969. Ecological factors and the distribution of Cladophora
glomerata in the Great Lakes. Am . Midland Naturalist 82:90-98.

Hutchinson, G. E. 1975. A Treatise on Limnology . Vol. III .
Limnological Botany . John Wiley and Sons. New York.

, and V. T. Bowen. 1947. A direct demonstration of the

phosphorus cycle in a small lake. Proc . Nat . Acad . Sci . 33:148-153.

, and ^. 1950. Limnological studies in Connecti-

cut. IX. A quantitative radiochemical study of the phosphorus cycle in
Linsley Pond. Ecology 31:194-203.

International Joint Commission. 1975. Cladophora in the Great Lakes.
IJC Regional Office, Windsor, Ont.

Lin, C. K. 1971. Availability of phosphorus for Cladophora growth in
Lake Michigan, pp. 39-43. In : Proc. 14th Conf. Great Lakes Res.,
Internat. Assoc. Great Lakes Res.

. 1977. Accumulation of water soluble phosphorus and hydrolysis

of polyphosphates by Cladophora glomerata . J^. Phycol . 13:46-51.
, and J. L. Blum. 1973. Adaptation of eutrophic conditions by

Lake Michigan algae. Tech. Rep. Wis. WRC 73-03. The University of
Wisconsin, Water Resources Center, Madison, Wisconsin.

, and . 1977. Recent invasion of a red alga

( Bangia atropurpurea) in Lake Michigan. sJ. Fish . Res .
Board Can. 34:2413-2416.

McRoy, C. P., and R. J. Barsdate. 1970. Phosphate absorption in eelgrass,
Limnol . Oceanogr . 15:6-13.

, , and M. Nebert. 1972. Phosphorus cycling in

eelgrass ( zostera marina L.) ecosystem. Limnol . Oceanogr . 17:58-67.

Neil, J. H. 1974. Cladophora in the Great Lakes. Prep, for IJC by
Limnos Ltd.

, and G. E. Owen. 1964. Distribution, environmental requirements

and significance of Cladophora in the Great Lakes, pp. 113-121. In: Proc.
Seventh Conf. Great Lakes Res., Internat. Assoc. Great Lakes Res.

117

Nichols, D. S., and D. K. Keeney. 1973. Nitrogen and phosphorus release from
decaying water milfoil. Hydrobiologia 42: 509-525 •

Peverly, J. H. , and J. Brittain. 1978. Effect of milfoil ( Myriophyllum

spicatum L.) on phosphorus movement between sediment and water. J^. Great
Lakes Res . 4:62-68.

Reimer, D. N., and S. J. Toth. 1968. A Survey of the Chemical Composition
of Aquatic Plants in New Jersey . N. J. Agr. Exp. Stat. Rutgers Univ.
Bull. 820.

Sculthorpe, C. D. 1967. The Biology of Aquatic Vascular Plants .
St. Martins Press. New York. 610 pp.

Taft, C. E., and W. J. Kishler. 1973. Cladophora as related to pollution

and eutrophication in western Lake Erie. Water Resources Center, The Ohio
State University, Columbus, Ohio.

Wezernak, C. T., D. R. Lyzenga, and F. C. Polcyn. 1974. Cladophora dis-

tribution in Lake Ontario. USEPA Report No. EPA-660-3-74-028. 76 pp.

118

CHAPTER VIII

CONCEPTUAL MODEL OF PHOSPHORUS CYCLING

Donald Scavia

Great Lakes Environmental Research Laboratory

National Oceanic and Atmospheric Administration

Ann Arbor, Michigan

119

INTRODUCTION

There presently exists a broad spectrum of models addressing phosphorus in
lake ecosystems • These models range from relatively simple, empirical models
of annual, lake-wide average total phosphorus concentration to more detailed
seasonal, spatially segmented models describing dynamic interactions among many
phosphorus components. Some models deal directly and solely with phosphorus
while others track phosphorus stoichiometrically with carbon or other biomass
variables. The motivation behind the different types of models also varies.
Models with simple physical segmentation and few ecosystem compartments have
been developed primarily for assisting water quality management, whereas other
models, which include detailed segmentation and many compartments, have been
developed generally for increasing our understanding of structure and function
within the ecosystem. The purpose of this chapter is to briefly outline
several categories of existing models and to construct a conceptual model
useful for synthesizing and analyzing existing information on phosphorus
cycling in lakes .

PREVIOUS MODELS
Management Models

The first category of models includes those designed around the phosphorus
loading concept. A thorough historical review and quantitative critical
valuation of these models is provided by Reckhow (1979). The development and
refinement of the mass-balance-based phosphorus model for lakes is attributed
to Vollenweider (1964, 1969) with significant alterations and improvements by
Dillon and Rigler (1974) and Chapra (1975, 1977). These models are simply
balances of phosphorus inflow, outflow, and sedimentation; the latter process
is used to calculate the retention coefficient (R) for phosphorus. Expressions
used for calculation of R have been examined statistically on several different
data sets (Dillon and Rigler 1975, Vollenweider 1975, Kirchner and Dillon 1975,
Chapra 1975, Larsen and Mercier 1975). A]. so, Reckhow (1977, 1979) and Walker
(1977) evaluated overall model accuracy for this set of models using
first-order error analysis. Modifications of the model have included sediment
interactions (Lorenzen et al. 1976) and separation of epilimnion/hypolimnion
and particulate/dissolved phosphorus compartments (Snodgrass and O'Melia 1975).

Mthough these models represent the only phosphorus-related models that
have undergone extensive statistical evaluation with respect to general
applicability across a wide range of lakes, they do not reveal much of the
structure or function of the ecosystems. This black-box, mass-balance approach
intentionally removes the effects of internal system dynamics. The models are
designed specifically to assess whole lake trophic status (phosphorus and
chlorophyll content) with respect to past,, present, and expected phosphorus
loads which reduces their effectiveness for examination of process-oriented
cycling of phosphorus.

121

Eutrophication Models

A second category of models falls between the two extremes outlined above
(management and ecosystem dynamics). These models, here termed eutrophication
models, are concerned with simulation and prediction of seasonal phytoplankton
biomass dynamics. This type of model, based on the early work of Riley et al.
(1949), was reintroduced in the early 1970s (e.g. Chen 1970, DiToro et al.
1971) and has had several significant alterations and applications (e.g.
Thomann et al . 1975, Chen and Orlob 1975, DiToro et al . 1975, DiToro and
Matystik 1980).

These models focus on one or two groups of phytoplankton and use
chlorophyll or dry weight as biomass indicators. A guiding principle behind
development and tests of these models is that ecosystem processes will be
included in the model only if they have significant impact on simulation and
prediction of these crude indicators of phytoplankton biomass. Consequently
processes such as phytoplankton production, respiration, and sinking;
zooplankton grazing, respiration, and excretion; and crude cycles of phosphorus
(two compartments) and nitrogen (three compartments) have been included.
Properties and processes such as internal storage of excess nutrients by
phytoplankton, minimum threshold of food for zooplankton grazing, multiclass
descriptions of phytoplankton and zooplankton, and food-supply control of
zooplankton respiration have not been included.

For phosphorus, the eutrophication models generally track the following
classes: 1) available phosphorus, 2) non-living unavailable phosphorus, and
3) living plankton phosphorus. The last class of phosphorus is usually
followed stoichiometrically with other biomass indicators (chlorophyll, carbon,
dry weight) . The models are not designed to follow details of the phosphorus
cycle. Model component definitions are often operational and tied to their
utilization by phytoplankton. Although this class of models has provided
important advances in the use of more mechanistic models in management of
aquatic resources, they too fall short of providing the detail and focus
necessary to explore phosphorus cycling quantitatively.

Ecosystem Models

This class of models, although fundamentally similar to the eutrophication
models, places emphasis on more detailed aspects of ecosystem dynamics. These
models are oriented toward synthesizing existing information, simulating known
conditions, and analyzing controls of system structure and function. Some
models of this type have been used for prediction, however their applicability
has not been tested as rigorously as the two other classes of models.

Ecosystem models emphasize more detailed aspects of plankton dynamics than
do eutrophication models. For example, seasonal succession (e.g. Lehman et al.
1975, Bierman 1976, Scavia 1980), internal stores of nutrients (Lehman et al.
1975, Bierman 1976), size-selective, minimum- threshold zooplankton grazing
(Scavia 1980), sediment-water column interactions (Larsen and Mercier 1975,
Scavia 1980), bacteria dynamics (Bloomfield 1975, DePinto 1979), zooplankton
vertical migration (Steele and Mullin 1977), and population structure (Steele
and Frost 1977) have been included in several ecosystem models. Scavia (1979)
provided a review of several detailed process components included in lake
ecosystem models.

122

Ecosystem models are collections of formalized theories of process
dynamics and are useful for describing and analyzing detailed aspects of the
phosphorus cycle. One shortcoming is that they presently emphasize phosphorus
dynamics only as they are directly related to plankton nutrition. The omission
of detailed dynamics for nonliving phosphorus components reflects both the
orientation of these models and the lack of information presently available on
those components.

Phosphorus-based Ecosystem Models

Recently, models focusing on phosphorus as a system tracer have been
developed. The models, most notably Lean (1973a), are based on radiotracer
experiments and generally assume linear, steady-state kinetics. A variety of
studies and experiments (Lean 1973a, b) have led to the following conceptual
model :

Colloidal

Seston

where XF represents phosphorus compounds of relatively low molecular weight
(250 Ml/^) , colloidal P represents relatively high molecular weight
(>5,000,00 MW) compounds, PO4 is dissolved orthophosphate, and seston is
phosphorus associated with particles retained on a membrane filter with 0.45 ym
pore size. Experimental work (Lean 1973a, 1973b; Lean and Rigler 1974, Lean
and Nalewajko 1976, Paerl and Lean 1976) also has led to the following
information: 1) Transfer from PO4 through XP to colloid requires the presence
of seston, indicating XP is excreted or otherwise released by plankton; 2) The
pathway XP -> Colloid -> PO4 is probably an exchange process of XP for PO4
associated with the colloids, PO4 is not adsorbed to the colloids; 3)
Time-course studies showed decomposition of XP to PO4 in 4 hr in a sterile
environiaent , indicating the chemical instability of XP (^^^250 MW) compounds;
4) Kinetics of exchange with seston P indicate at least two seston compartments
- a fast exchanger and a slow exchanger.

Phosphorus-based models provide important first steps in identifying
significant nonliving phosphorus components and major pathways of phosphorus
flow under steady-state, experimental conditions. However, because of their
steady-state orientation, they do not provide theory for, nor tests of,
hypotheses regarding seasonal, dynamic interactions among phosphorus
components. Also, they are not oriented toward food-web dynamics and consider
only radioactive labeled compounds over a relatively short time interval.

The above brief descriptions of extant modeling capabilities indicate
that, while each class of models has attributes specific for its intended use,
they all are incomplete with regard to examination of detailed phosphorus
cycling for two reasons. Firstly, they generally do not follow many of the
important components of phosphorus. The first class follows only total
phosphorus and the other two classes follow only components directly related to

123

phytoplankton nutrition. The keys to many unresolved questions relating to
utilization, cycling, and biological availability of phosphorus are likely in
phosphorus components of the nonliving, non- or t ho phosphate phosphorus pools.
Secondly, those models generally do not address many of the current theories
regarding controls of internal nutrient cycling. Only by including those
theories (e.g., the effects of animal migration and patchiness in general,
abiotic/biotic and phytoplankton/ bacteria competition for nutrient ions, and
sedimentary and resuspended material) , can their significance in the larger
ecosystem context be examined.

It is for these reasons that a conceptual model identifying many of the
missing components and dynamics is proposed below.

CONCEPTUAL MODEL OF PHOSPHORUS CYCLING IN THE GREAT LAKES

This conceptualization is not intended to be implemented computationally
in its entirety, but rather is an attempt to recognize potentially important
aspects of the phosphorus cycle. Subsequent definitions of appropriate time
and space scales and consideration of the physics will assist in identification
of important and consistent (for those scales) components and processes.

The material that follows is based, in large part, on information drawn
from the preceding chapters and represents a synthesis of the material reviewed
and ideas proposed into a somewhat complete whole-system framework. This
framework serves as a background against which the significance of individual
components and processes can be judged.

Phosphorus Transport

Figure 1 is a representation of major physical and migratory processes
that are superimposed all over biological and chemical in situ transformations.
Boyce (1974) and Lick (1979) reviewed aspects of large- and small-scale
circulation, vertical and horizontal dispersion, and waves and resuspension as
they relate to biological and chemical properties and transport of
contaminants. Vertical and horizontal migration of zooplankton, fish, and
benthic animals can, at times, be significant modes of transport. This is
particularly true when dominant physiological processes depend on position in
the migratory patterns (see Bowers, Chapter IV). Sinking ensures a long-term
net downward transport of particulate material. Experience has indicated that,
for large, spatially (horizontal and vertical) averaged Great Lakes models,
phytoplankton sinking is not a particularly important mass flux process on
short time scales (<1 year). On longer time scales, sinking is the dominant
process. Sinking of nonliving particulate material may be significant on both
time scales. The significance is tied closely to the scavaging effect of
sorption onto these particles.

All of these processes interact and impact on biological and chemical
properties in lakes; their relative importance is keyed to time and space
scales that are under investigation (see Harris 1980). To compare the
influence of these transport phenomena to rates of in situ transformations.
Table 1 was constructed. For processes like these, only the spatial scales of
interest need be defined, for time scales are dictated by the process and space
scale. In the first column in Table 1 maximuun expected rates of transport are

124

Sources and Transport Processes

Fig. 1. Phosphorus sources and transport.

TABLE 1. Time constants for transportation phenomena.

Maximum
Transport Rates

10 m day"-^
10 m day*"^

10,000 m day"l

1,000 m day""^

100 m day"^

10,000 m day""^

10 m day""-*-

1 m^day""-*-

106in2day'"^

Upwelling

Downwelling

Large-scale circulation
(horizontal)

Small-scale circulation
(vertical)
(e.g., Langmuir)

Migration: zoo (vertical)

Migration: fish (horizontal)

Sinking

Dispersion : vertical

horizontal

Upper Limit
Rate Constant

10 day-"!

10 day"i

10 day""l

1,000 day""^

100 day"l

10 day""l

10 day""^

1 day
1 day

-1

-1

Minimum vertical scale = 1 m
Minimum horizontal scale = 1000 m

125

given. Assuming minimal spatial scales of interest of 1 m (vertically) and 1
km (horizontally) then upper limit, reciprocal residence times (column 2) can
be calculated. These time constants quantify the maximum effect of the
transport processes on the given spatial scales and can be used to identify
those processes that "compete" with biological and chemical processes
(discussed below) on these scales. Other space scales can (and should) be
considered. Also, these assumed space scales will not necessarily resolve all
transport processes .

Phosphorus Components in the Water Colximn

At any particular location within the lake, many biological and chemical
processes are acting simultaneously to transform phosphorus among its many
chemical forms. A representation of those forms and probable important
pathways of phosphorus flux in the Great Lakes are illustrated in Figure 2.

Living Components

The right half of the figure is dominated by living components including
food web and fecal pathways among phy to plankton, bacteria, protozoa, rotifers,
crustaceans, and fish. Although difficult to illustrate on the figure, some of
the most important aspects of the control of phosphorus dynamics from the top
of the food web are keyed to predatory effects on community structure. For
example, given increasing metabolic rates with decreasing animal size, the
size-dependent nature of fish predation on the zooplankton community becomes
increasingly important (see Kitchell, Chapter V). To investigate these and
similar effects, several functional groups and perhaps life-stage categories
must be considered throughout the food web. Suggested criteria for separation
of functional groups are included in Table 2.

Another property that has been included previously in some models, and
apparently must be included, is internal pools of readily exchangeable
phosphorus within bacteria and phytoplankton cells. While the kinetics of
exchange among external, internal, and biomass pools of phosphorus must be
considered carefully in the context of specified time scales, there is no
question that the consequence of interaction among these pools - variable cell
nutrient stoichiometry - must be included (see Lehman, Chapter IV).

Nonliving Components

The left half of Figure 2 represents the nonliving matrix of phosphorus
components. Here truly soluble components exist "freely dissolved" or sorbed
to particulate material. Sorption processes may serve to buffer the more
transient processes by providing a reservoir of material. This reservoir is
built up during periods of high concentration of dissolved material and
subsequently released during periods of lower concentration.

Dissolved inorganic components of this nonliving matrix can be separated
out easily on functional, if not analytical, grounds. For example, dissolved
or tho phosphate (PO4) fills the special role of being the end product of most

126

Carnivore
Zooplankton*

Fig. 2. Phosphorus flow diagram indicating major components and
pathways within the water column. * indicates several functional
groups or life stages within each component. Numbers represent
processes identified in Table 4.

degradation processes, as well as being the only form of phosphorus immediately
available for assimilation by algae. Although conventional analytical methods
can only approximate the concentration of PO4 (see Tarapchak, Chapter I;
Gardner and Eadie, Chapter II), it is important to understand and attempt to
describe the flux of phosphorus through this important pool. Inorganic
condensed phosphates may or may not play an important role in the phosphorus
cycle. These compounds (e.g., the polyphosphates) are the principal storage
products of surplus phosphorus in algae, but their stability and importance
outside the cell are not well known. This conceptualization (Figure 2)
recognizes the potential importance of this group of compounds. A suite of
other dissolved inorganic compounds (e.g., inorganic complexes) may also exist,
although their relative abundance and importance are unknown.

Dissolved organic phosphorus compounds are probably the least well
understood group. Condensed organic phosphates (e.g., ATP, ADP) are key
compounds in metabolic and energetic processes within organisms, but their
relative stability and importance as compounds outside organisms are not
clearly understood. Organic compounds, other than these condensed forms, fall
into a mesh of overlapping categories (e.g., humic acids, organic complexes,
phytic acids, nucleotides). The role of this group, or more accurately the

127

TABLE 2. Criteria for disaggregation of functional groups.

Particulate Organics (+ other particulates)

— size

— lability

— sorption capacity

Dissolved Phosphorus

— inorganic/ organic

— lability

— sorption characteristics

Bacteria

— food sources

— planktonic

— "attached"

Phytoplankton

— size

— resource requirements (Si, P, N)

— kinetic response capability

Zooplankton

— size

— food selection

— kinetic response capability

— life history

Fish

— food selection

— life history

roles of components within this group, has not been studied extensively. At
this stage only speculation can be used to identify the important components.
Perhaps the best approach to conceptualize important components is to visualize
the compounds on a "degradability" or "labile-refractory" continuum, which may
or may not be related to component size (or molecular weight). This is
discussed further below with particulate organic P.

The colloidal pool is identified on Figure 2 to emphasize the fact that
this pool (mainly a size-dependent definition) can be important. Because this
pool is defined only by size characteristics, its functional characteristics
include a wide spectrum of processes and attributes (both inorganic and
organic) and it therefore requires the same scrutiny relating to lability as
the truly dissolved components.

Particulate components of the nonliving pool generally can be divided into
inorganic and organic classes, although organic coatings on inorganic particles
and inorganic "intrusions" (e.g., silica frustules in fecal pellets) make even

128

this distinction difficult. The role of particulates falls into two categories
in terms of controls of phosphorus cycling: 1) a site for sorption of
dissolved components, 2) a source or sink for dissolved components from
phosphorus incorporated into the particulate material. Examples of the
sorption role are sorption onto CaC03 particles, clays, and metal precipitates
[e.g. Fe(0H)3]. Particulate inorganic phosphorus compounds are generally
considered to be not important in the water column. Examples of the second
role are phosphorus compounds found in dead algae fragments and zooplankton
fecal material. Algal fragments and fecal material can also serve as sorption
sites .

Like the dissolved components, particulate organic phosphorus components
represent spectra of size and lability. Although particulate and dissolved
components are represented on Figure 2 in discrete classes. Figure 3 better
represents this nonliving matrix. Entries in the matrix represent only what
can be implied from the paucity of available information: that a large fraction
of naturally occurring phosphorus compounds in this nonliving matrix must be
classified presently as amorphous, and that the limits of lability indicated on
Figure 3 for many of the constituents are not well known. As more information
becomes available to fill out the matrix, not only may logical functional
groups of compounds be identified, but also potential classification of
constituents in water samples of unknown composition may be possible.

Typical Pool Sizes

Table 3 lists typical summer pool sizes. For many of the components (e.g.
condensed P, colloidal P, protozoa) very little or no information is available
for the Great Lakes. Even where data are available, they are general and
crude. Orthophosphorus is probably less than 0.1 yg P L"^ in many lakes since
SRP measurements (likely an overestimate; see Tarapchak, Chapter I) in Lake
Ontario, for example, are ca. 1.0 yg P L""^. SUP, which estimates the nonliving
matrix (Figure 3) in the less-than-0.45 ym size range, in Lake Ontario is
typically on the order of 10 yg P L""^. Phytoplankton P, estimated from direct
counts, carbon per cell, and Redfield's ratio, is ca. 5 yg P L"^ in Lake
Ontario. The nonliving >0.45 ym size range (PP), approximated by total P
(TP=25 yg P L""^) - total dissolved P(TDP) - phytoplankton P, contains about
10 \xg P L"^ in Lake Ontario. Phytoplankton and detrital organic carbon
concentrations are comparable in the summer epilimnion of Lake Ontario.
Assuming (as a maximum for detritus) the same is true in terms of phosphorus,
TP-TDP-phytoplankton P-Detrital P = 5 yg P L"^ provides an estimate of
phosphorus associated with inorganic particles, leaving 5 yg P L~^ associated
with organics. Since little epilimnetic P is truly particulate inorganic, the
former component is likely to be phosphorus sorbed to inorganic particles.

Phosphorus concentrations associated with herbivorous, omnivorous, and
carnivorous crustaceans, approximated by simulation and by dry weight
observations converted stoichiometrically, are typically 1.0, 1.0, and
0.1 yg P L^l, respectively (see Scavia 1980). Rotifer concentrations in
nearshore Lake Ontario are on the order of 1000-1500 ind L"^ (McNaught et al.
1973). Using dry weight estimates from Nalepa (1972) and a phosphorus content
of about 1%, rotifer phosphorus may be ca. 0.1 yg P L""^.

Bacterioplankton typically number 10^ cells mL"^ in lakes and oceans (e.g.
Wiebe and Pomeroy 1972), and one study (Barsdate et^ al. 1974) of a tundra pond

129

"Stability" (Half-life, seconds)
(Decomposition, Digestion, Reaction, Uptake, etc.)

102

103

1

104 1

05

106

107

108

10S

)

SIZE

(MW) (ytim)

1

MINUTES

HOURS

DAYS

WEEKS

MONTHS

YEARS

DECADES

10^ 10-5-

Q

5

1

hi

|P0.|

|Sugar-P04 |

Poly-F

' Inositol Hexaphosphate]

103 10-4-

1 Nucliotides(ATP, ADR)

jPhospholipids |

■ ■* ""

105 10-3-

Humics

i
<

c

J
c

■ MB
« Ml MM ■■

Nucleic A

^:^« 1

107 10 2-

Algal Fragments

10-1-

- o
-J
-J

Diffuse Feces

100-

c

c

J)

101 -

Intact Algal Cells

Dead Animal Residues

10^ -

3
O
P
(X.

- ^

Fecal Pellets

103-

Mineral
(e.g. Apatite)

104-

1

1

Fig. 3. Size-stability matrix of nonliving phosphorus components indicating
approximate location of identifiable fractions.

ecosytem measured 2.6x10""^ \xg P cell"^ for bacteria. These numbers indicate a
possibility of ca. 2.6 yg P L""^ due to bacteria alone. Concentrations of
bacteria attached to particles are perhaps one-tenth of free swimming bacteria
(see Moll and Brahce, Chapter III).

In Situ Transformations

Important phosphorus components are illustrated in Figures 2 and 3 and
typical pool sizes are listed in Table 3. Important processes contributing to
transport of those components are illustrated in Figure 1 and typical rates are

130

TABLE 3. Typical pool sizes.

Pool

Concentration (yg P L"^

Dissolved orthophosphate (<SRP)

Nonliving (<.45 yin, SUP)

Nonliving (>.45 ytn, organic)

Sorbed to Inorganic Particles

Bacterioplankton

Attached Bacteria

Phytoplankton

Herbivore crustaceans

Omnivore crustaceans

Carnivore crustaceans

Rotifers

Planktivore fish

Piscivore fish

Total P

<0.1
10-0

5-0

5.0

1.0-10.0

0.1-1.0

5.0

1.0

1.0

0.1

0.1

0.01-1.0

0.001-0.1
25

listed in Table 1. Important biological and chemical processes that cause
transformations of the components resulting in fluxes among the boxes in Figure
2 and, in effect competing with the processes listed on Table 1, are
illustrated as numbered arrows on Figure 2. The processes are listed by those
numbers in Table 4 where typical, phosphorus-specific rate constants are given.
These rates are by no means a synthesis of all existing information. They
represent typical ranges of values accumulated through a cursory investigation
of the literature. These rate constants multiplied by the expected pool sizes
(Table 3) indicate approximate fluxes among the compartments.

Examination of fluxes presents a crude approximation of the relative
importance of various process/ pool interactions in controlling the cycling of
phosphorus. It also illustrates relative time scales of consideration for
processes. For example, on a 1-day time scale bacterial uptake may appear
always in equilibrium with external P concentrations, whereas animal
nonpredatory death may appear not to occur at all.

Simultaneous consideration of physical, biological, and chemical processes
affecting the phosphorus cycling will require careful matching and selection of
time and space scales. Biological and chemical processes synthesized into
general categories (Table 5) should be compared with the transportation process
rates in the second column of Table 1. The rate constants are compared
graphically, assuming minimum transport space scales of 1 m vertically and 1 km

131

TABLE 4. Detailed rate constants and estimates of fltix.

Rate Constant

Pool Size

Flux

No. Process

(day-1)

(ygP L-1) (ygP

L-lday"l)

(1) Sorption

PO4 to seston

10

<.l

<1.0

PO4 to Hydroxides

100

<.l

<10

(2) Death

Bacteria

.001

1-10

.001-. 01

Phytoplankton

.01-. 1

1-10

.01-1.0

Herbivore crustaceans

.01

1.0

.01

Omnivore crustaceans

.01

1.0

.01

Carnivore crustaceans

.01

.1

.001

Planktivore fish

.01

.Ol-.l

.0001-. 001

Piscavore fish

.01

.001-. 01

.00001-. 0001

Rotifers

.01

0.1

.001

Protozoa

.01

?

7

(3) Excretion/Egestion

Bacteria

10

1-10

10-100

Phyto (to DOP)

0.1

1-10

0.1-1.0

Herbivore crustaceans

.001-10.

1.0

.001-10.0

(to PO4)

Omnivore crustaceans

.001-1.0

1.0

.001-1.0

(to PO4)

Carnivore crustaceans

.001-1.0

0.1

.0001-. 1

(to PO4)

Rotifers

7

7

?

Planktivore fish

.001

.Ol-.l

.10-5-10-^

Piscavore fish

.001

.001-. 01

10-6-10-5

Protozoa

?

?

7

"Decay"

ORG - Inorg

.01

15(P0P-H)0P)

.15

(4) Uptake

Phytoplankton

10

1-10

10-100

Bacteria

100

1-10

100-1000

(5) Enzymatic release

(alk. phos. on DOP)

0.1-1.0

10

1-10

(6) Photolysis (of DOP)

1-10

10

10-100

(7) Hydrolysis

of Cond P

•Ol-.l

7

?

of org P

1-10

10

1000

(continued)

132

TABLE 4. (continued).

Rate Constant

Pool Size

Flux

No. Process

(day-1)

(ygP L-1)

(ygP L-lday-1-)

(9)

Feeding

Protozoa

?

7

?

Herbivore crustaceans

0.1-1.0

1.0

.1-1.0

Omnivore crustaceans

0.1-1.0

1.0

.1-1.0

Carnivore crustaceans

0.1-1.0

0.1

.Ol-.l

Planktivore fish

0.01-0.1

.Ol-.l

.0001-. 01

Piscavore fish

0.01-0.1

.001-. 01

.00001-. 001

Rotifers

0.1-1.

0.1

.Ol-.l

(10)

Growth

Phytoplankton

1

1-10

1-10

(on internal P)

Bacteria

10

1-10

10-100

(on internal P)

TABLE 5. List of general biological and chemical processes.

No.

1
2
3
4
5
6
7
8
9
10
11

Process

Rate Constant (day~^)

Sorption

Death

Excretion/ egestion

Uptake

Enzymatic release

Photolysis

Hydrolysis

Precipitation reactions

Feeding

Growth (on internal P)

Microbial colonization

10^-102

10-3-10-1

lO-3-iol

10^-102

lO-l-ioO

loO-ioi

lO-2-iol

7
lO-2-ioO

loO-ioi

133

horizontally, in Figure 4. From this display it becomes clear that careful
attention must be given to the interactions of physical, biological, and
chemical processes on varying time and space scales. This is true for both
modeling and experimental considerations (see Harris 1980).

Sediment/Water Interface

To this point only processes and components affecting phosphorus cycling
in the water coluann have been discussed. Transport across the sediment/ water
interface, also an important process, is affected by many properties and
processes interacting within the sediment. The most important factors are
identified in Figure 5.

The sediment region is composed of particulate material, pore water
contituents, and permanent and migratory organisms. Transformation of
phosphorus species, translocation within the sediment, and translocation into
and out of the sediment are processes labeled on Figure 5 and identified in
Table 6. Along with each process or property in Table 6 is a suggested list of
functions controlling that process and, where available, estimates of flux
rates. Presently, the qualitative controls of phosphorus exchange are
beginning to emerge, however very little is known of the quantitative and
mechanistic nature of these controls.

DISCUSSION

The material in this chapter represents a synthesis of information from
the previous chapters as well as other literature sources. An attempt is made
to outline important components (Figures 2 and 5) of the phosphorus cycle and

10

10 ^

lO-i

Turnover Rate (day"'')

100 101

102

103

104

-Enzyme Release -

I

^ Upwelling ^
Downwelling

^ Fish ^
Migration

I

_ Large Scale _^
" Circulation

^Zooplankton^^
Migration

_ Small Scale _
" Circulation "

— Nonpredatory Death -

.Vertical and ^

Horizontal Mixing

- Growth -

-Excretion and Egestion-

- Feeding -

-Hydrolysis-

~ Sinking — ►

— Uptake -
-Sorption—

-<- Photolysis-^

Fig. 4. Comparison of physical and biological/ chemical process time scales.
Physical processes assumed important on 1 m vertical and 1 km horizontal
space scales-

134

TABLE 6.. Processes at the sediment-water interface.

No.

Processes

Rates

Flux

1 Deposition (particle size,

composition)

2 Resuspension (waves;

currents ; reworked ,
uncompacted sediments
from processes
(4) and (5))

3 Animal migration

(translocated pro-
cesses ; meiof auna)

lO-^-lO-^ gPni-2 day~l

lO^-lol m hr-1

Pontoporeia

( "eddy-diff users"
of particulates;
see (2))

Tubificids (conveyor
transport of solids
and water from
2-5 cm to top;
see (3))

Chironomids (water
pumpers mainly,
from 2-5 cm)

7 Pore water and inter-

face diffusion

8 Particulate - pore water

exchange (many processes
similar to heterotrophic
water column; chemical
equilibrium - sorption;
biological egestion excre-
tion, feeding macro-,
meio-, micro-benthos)

10-5.-10-6 jq2 4ay"l

10""2-io-l gPm""2 day""^

See Figure 3 and Table 4

135

SEDIMENT ; • •

PARTICULATES

cm

8

PORE WATER

Fig. 5. Conceptualization of major processes acting at the sediment-
water interface. Numbers represent processes identified in Table 6.

to suggest a logical basis (Figure 3) for identifying other important
components, especially within nonliving phosphorus pools. Pool sizes typical
for the Great Lakes (especially Lake Ontario) are given (Table 3) for
components where data are available.

Important pathways of phosphorus flows among components are indicated
(Figures 2 and 5) and flux estimates are given (Tables 4, 5, and 6). Although
the flux estimates are crude, they can be useful for examining the relative
importance of the pathways. The true impact of each pathway also is determined
somewhat by time scales under consideration. While the pathways are quantified
somewhat by examining ranges of flux rates, no attempt was made to describe any
controls of processes identified as pathways. The next step in development of
pertinent information on phosphorus cycling will be to define and refine
equations that approximate those processes and their functional controls.

These transformation processes are often considered intrinsic properties
when viewed in light of water transport, i.e., reactions within a parcel of
water are not affected by the dimension of that parcel. This is only true,
however, for a homogeneous environment. Harris (1980) discusses some important
consequences of ignoring spatial heterogeneity in phytoplankton populations.
It appears that averaging a specific process (say phytoplankton uptake) over a
space scale several times the size of a phytoplankton patch will result in a
rate quite different from one within the patch. Consequently, measurement and
simulation of biological and chemical processes may be space-scale dependent
and therefore tied intimately to transport processes.

Some important transport processes are illustrated in Figure 1 and typical
rates identified in Table 1. Assuming minimum spatial scales of 1 m vertically
and 1 km horizontally (the scales at which much of the presently available
biological and chemical rate information is pertinent) the effects of transport
versus in situ transformation are compared in Figure 4. It is through
consideration of both transformation and transport processes at proper time and
space scales that the relative importance of the processes will emerge.

136

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Bierman, V. J., Jr. 1976. Mathematical model of the selective enrichment of
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