sources and bioavailability of phosphorus fractions in freshwaters: a british perspective

$: Sources and bioavailability of phosphorus fractions in freshwaters: a British perspective$
Biol. Rev. (2001) 76, pp. 27–64 Printed in the United Kingdom # Cambridge Philosophical Society 27

Sources and bioavailability of phosphorus

fractions in freshwaters : a British perspective

C. S. REYNOLDS and P. S. DAVIES"

CEH Windermere Laboratory, The Ferry House, GB-LA22 0LP Ambleside, Cumbria, UK

(Received 2 July 1999; revised 11 July 2000; accepted 24 July 2000)

ABSTRACT

This paper seeks a perspective on the forms of phosphorus which promote aquatic eutrophication, with theparticular quest of establishing their sources. A short background traces the development of understandingof nutrient enrichment and the suppositions about the relative contributions of agriculture, sewage anddetergent residues. Most aquatic systems, and their primary producers, are naturally deficient inbiologically-available phosphorus. Aquatic plants have evolved very efficient phosphorus uptakemechanisms. The biomass responses to an increase in the supply of phosphorus are stoichiometricallypredictable. The most bioavailable forms of phosphorus are in solution, as orthophosphate ions, or arereadily soluble or elutable from loose combinations. Ready bioavailability coincides well with what ismeasurable as molybdate-reactive (MRP) or soluble-reactive phosphorus (SRP). Most other forms,including phosphates of the alkaline earth metals, aluminium and iron are scarcely available at all.Orthophosphate ions sorbed to metal oxides and hydroxides are normally not biologically available either,except through weak dissociation (‘desorption’). The production of alkaline phosphatase provides organismswith an additional mechanism for accelerating the sequestration of phosphate from organic compounds.Bioavailable phosphate is liberated when redox- or alkali-sensitive metal hydroxides dissolve but theseprocesses are minor contributors to the biological responses to nutrient enrichment.

Most of the familiar eutrophication is attributable to the widespread application of secondary sewagetreatment methods to the wastes emanating from a burgeoning and increasingly urbanised humanpopulation. The use of polyphosphate-based detergents, now in decline, has contributed to the problem. Inaquatic systems, the additional phosphorus raises the biological supportive capacity, sometimes to thecapacity of the next limiting factor (carbon, light, hydraulic retention or of another nutrient). At highorthophosphate loadings, the straight stoichiometric yield relationship between biomass yield and phosphorusavailability is lost.

Movements of phosphorus and its recycling within aquatic systems do not prevent the slow gravitation ofphosphorus to the bottom substrata. The phosphorus retentivity of sediments depends upon their chemicalcomposition. While oxide-hydroxide binding capacity in the surface sediments persists, they act as a sink forphosphorus and a control on further cycling. Iron-rich and clay-rich sediments perform best in theseconditions ; calcareous sediments least so. Eutrophication may lead to the exhaustion of sediment P-bindingcapacity. Non-sorbed phosphate is readily recyclable if primary producers have access to it. Recycling is mostrapid in shallow waters (where sediment disturbance, by flow, by wind action and through bioturbation, isfrequent) and least in deep ventilated sediments.

The contributions of phosphorus from catchments are assessed. The slow rate of weathering of (mostlyapatitic) minerals, the role of chemical binding in soils and the incorporation and retentivity by forestedterrestrial ecosystems each contribute to the minimisation of phosphorus leakage to drainage waters.Palaeolimnological and experimental evidence confirms that clearance of land and ploughing its surfaceweakens the phosphorus retentivity of catchments. The phosphorus transferred from arable land to drainageremains dominated by sorbed fractions which are scarcely bioavailable. Some forms of intensive marketgardening or concentrated stock rearing may mobilise phosphates to drainage but it is deduced that drainagefrom agricultural land is not commonly a major source of readily bioavailable phosphorus in water. Careful

" Present Address : Department of Geography, University of Reading, GB-RG6 6AB Reading, UK.

28 C. S. Reynolds and P. S. Davies

budgeting of the phosphates in run-off from over-fertilised soils may nevertheless show that a proportionatelysmall loss of bioavailable phosphorus can still be highly significant in promoting aquatic plant production.The bioavailable-phosphorus (BAP) load achieving the OECD threshold of lake eutrophy (35 mg P m−$) iscalculated to be equivalent to a terrestrial loss rate of approximately 17.5 kg BAP km−# year−"), or only1–2% of a typical fertiliser application. The output is shown to be comparable with the P yield fromsecondary treatment of the sewage produced by a resident population of 30–44 persons km−#. With tertiarytreatment, the equivalence is with approximately 200 persons km−#.

CONTENTS

I. Introduction ............................................................................................................................ 28(1) Objectives of the review .................................................................................................... 28(2) Phosphorus in fresh waters................................................................................................ 29(3) Cultural eutrophication..................................................................................................... 30(4) Magnitude of phosphorus loads ........................................................................................ 31

II. Chemical speciation of phosphorus in natural waters ............................................................. 32(1) Sources .............................................................................................................................. 32(2) Phosphorus fractionation and biological availability ........................................................ 33(3) Relative distribution of phosphorus fractions in natural waters........................................ 35

III. The assimilatory value of aquatic phosphorus fractions .......................................................... 35(1) The Vollenweider-OECD model....................................................................................... 35(2) Cell stoichiometry ............................................................................................................. 36(3) Phosphorus uptake ............................................................................................................ 38(4) Estimating the bioavailable fraction ................................................................................. 39

IV. Transformations and translocations of phosphorus in aquatic systems.................................... 39(1) Proximal fates of phosphorus loads in aquatic systems ..................................................... 39(2) Processing the bioavailable fractions ................................................................................. 40(3) The influence of macrophytes ........................................................................................... 41(4) Phosphorus diagenesis in bottom sediments ...................................................................... 42(5) Phosphorus-binding capacity of sediments ........................................................................ 42(6) Redox- and pH-sensitive transformations ......................................................................... 43(7) Whole-lake phosphorus budgets........................................................................................ 44(8) Some case studies .............................................................................................................. 45(9) Internal phosphorus recycling: the emerging overview .................................................... 46

V. Catchment sources of phosphorus............................................................................................ 47(1) Phosphorus in soils ............................................................................................................ 47(2) Fertiliser application strategies.......................................................................................... 48(3) Other catchment phosphorus sources................................................................................ 49(4) Transfer pathways from soils to surface waters................................................................. 49(5) Phosphorus transfers at the catchment scale ..................................................................... 51

VI. Discussion ................................................................................................................................ 57(1) Phosphorus fluxes to water and their impacts .................................................................. 57(2) The singularity of variable bioavailability ........................................................................ 57(3) The singularity of P transfer from land to water .............................................................. 58(4) The singularity of aquatic phosphorus processing............................................................. 59(5) Yield equivalence of phosphorus sources........................................................................... 60(6) Epilogue ............................................................................................................................ 61

VII. Conclusions .............................................................................................................................. 62VIII. Acknowledgements .................................................................................................................. 63

IX. References................................................................................................................................ 63

I. INTRODUCTION

(1) Objectives of the review

The eutrophication of freshwater ecosystems, whichis generally understood to refer to enrichment of

waters by inorganic plant nutrients and the increasedproduction of algae it can stimulate (OECD, 1982),has been much investigated. However, the practicalproblems of its abatement are still with us and theapproaches to their diagnosis and remediation

29Sources and bioavailability of phosphorus fractions in freshwaters: a British perspective

continue to be contentious. There is general agree-ment that increased loadings of phosphorus (P) todrainage have played a central role in the accel-erated eutrophication of rivers, lakes and coastalwaters. It is also widely recognised that the extraloads have emanated from several sources, whichinclude industry, sewage treatment or enhancedagricultural production. Each of these aspects hasbeen intensively investigated and reported withinthe appropriate specialist scientific literature in, forinstance, soil chemistry and biochemistry, sanitationand waste-water treatment and aquatic biology.There is a recognition that phosphorus reaches waterin many chemical forms but there is much lessagreement about which of these are biologicallyavailable (in the sense of being usable by algae andplants) and about how they should be controlled.

The importance of these topics is reflected in thenumber of reviews and authoritative compilationsthat are currently available. Some of these, indeed,have proved invaluable to us in the preparation ofthe present overview and they are extensivelyquoted. Our intent here is to relate, possibly for thefirst time, the symptomatic growth of algae to thesources and chemical accessibility of the forms ofphosphorus delivered to water, with particularreference to those generated in agriculture. Thereview is not concerned about whether there aresignificant transfers from agricultural catchmentsbut about how much might be stimulatory to algalgrowth compared to that of other anthropogenicallyenhanced sources of nutrient.

Mindful of the political interest in these questions,not least in identifying the proximal liability forremediation, we wish to emphasise our objectivity.In drafting this essay, we take the view that there isa collective societal responsibility for what amountsto the enormous anthropogenic acceleration ofplanetary phosphorus fluxes. A quadrupling of anincreasingly urbanised human population duringthe last century has necessitated improved sanitarystandards of sewage treatment. Generally, thetechniques for the rapid oxidation of the organiccarbon also bring about the mineralisation of otherorganic assimilates in biogenic wastes, includingthose yielding simple compounds of phosphorus. Thehuge expansion in the human population has beenvitally supported by a commensurate increase in theagricultural production of foodstuffs. This ‘greenrevolution’ has involved both an areal expansion inthe global land holdings husbanded for food pro-duction, subject to all the existing constraints ofgeological and climatic suitability and the practical-

ities of irrigation, and a great increase in the intensityof area-specific yields. Neither could have beenachieved without the widespread application ofinorganic fertilisers to offset the natural deficienciesin the mobility of phosphorus in most soils. Thefundamental truth is that modern agriculture andthe socio-economic structures that it sustains arefounded on the enhanced distribution, chemicalmobility and biotic transfers of phosphorus andnitrogen.

Whether the elements thus liberated are removeddirectly to water courses or indirectly through thehuman food chain and its oxidised wastes becomesreally rather academic. If, as has been estimated(Howarth et al., 1995), the current input of phos-phorus from the land to the sea has trebled relativeto the pre-agricultural period and that the enhancedfertility of lakes, rivers and coastal waters is directlyresponsible for the damage to aquatic ecosystems,then the relevant questions should really be aboutthe scale and impact of phosphorus fluxes to aquaticsystems. How much phosphorus is shed to drainagewaters, in what fractions and with what bioavail-ability? In what transformations does this bioactivityresult? How much is available for re-use and for howlong do natural systems continue to do so? Theanswers are also vitally important to the design ofstrategies to control and manage eutrophicationsuccessfully.

The article is organised along the following lines.First, a background is assembled to provide anoverview of what is often projected as a simple cause-effect sequence that increasing quantities of plantnutrients in water support more algae. The relation-ships are shown to be very complex. This complexityowes, in part, to the environmental geochemistry ofphosphorus in natural waters, which is addressed inthe second section. The differing bioavailabilities ofthe phosphorus fractions are reviewed in the nexttwo sections, taking into account the use, trans-formations, re-use and fates of the various fractions.A further section considers the catchment sources ofphosphorus, with special reference to the diffusesources and the pathways of their removal to water.The final section attempts to overview the generalimpacts of anthropogenic phosphorus sources onaquatic systems as a function of their bioavail-abilities.

(2) Phosphorus in fresh waters

The scientific study of lakes and rivers is a relativelyyoung discipline but its early practitioners were soon


Table 1. Phosphorus fractions in water and their nomenclature

Phase: reactivity Abbreviation Sensitivity Bioavailable?

Dissolved P DP Orthophosphate Readilyin solution

Soluble, SRP, DPcolloid-molybdate-reactive P MRP bound P

Particulate P PP ConditionallyH

#O-extractable P IMRP Interstitial MRP

NH%Cl-extractable P NH

%Cl-P Exchangeable P

Citrate-dithionite- Na#S#O

%-P Ferric-bound P

extractable PNaOH-reactive P NaOH-rP Al-bound P

Non-NaOH-reactive PP NaOH-nrP ScarcelyHCl-reactive P HCl-rP Apatite ; P co-precipitated

with CaCO$

Residual P resP

HCLO%-digestible P TP Various

(Total P)

able to recognise the important linkages of aquaticecosystems to the geology, geomorphology andhydrology of their terrestrial catchments. Lakes weresubdivided according to differences in communitymetabolism. Eutrophic (‘well-feeding’) lakes aredistinguished by having higher concentrations ormore persistent supplies of plant nutrients, whichsupport greater intensities of primary production(including of algae) and maintain populations ofpercoid and cyprinoid fish. Oligotrophic (‘ few-feeding’) lakes are poor in one or more essentialnutrients, typically limpid, unproductive and sup-porting sparse (though prized) populations of salmo-noid fish species. Moreover, the nutrients whosesupply distinguishes the productivity of either kindof lake were known to be carbon, nitrogen andphosphorus, even before some of them could bereliably analysed.

At the same time, it was becoming apparent thatmany fresh waters were becoming more fertile andmore productive, with oligotrophic waters comingincreasingly to resemble eutrophic ones (Gibson,1997). The causes of change were soon attributed toenrichment by nitrogen and, especially, phosphorus.In many instances, the changes could be explainedexclusively by the additional supply of phosphorus:a powerful tradition has been developed, throughthe work of, inter alia, Sakamoto (1966), Dillon &Rigler (1974) and Schindler (1977), that phosphorusis the principal controlling variable in lake ecology.The view is encapsulated in the models developed for

the Organisation of Economic Co-operation andDevelopment (OECD) by R. A. Vollenweider(1968) and co-workers. In its final form (Vollen-weider & Kerekes, 1980), the regression remains themost powerful statement of the high-order linkagebetween fresh waters and the supply of nutrients.

(3) Cultural eutrophication

The extra nutrients in lakes, rivers and coastalwaters and their effects are regarded as detrimentaland their role in promoting ‘cultural ’ eutrophicationneeded to be quantified. Whereas particular indus-trial activities were readily identifiable as sources ofenrichment, the generality of the problem nurtureda supposition that the principal cause was theenhancement of agricultural production and theintensified use of inorganic fertilisers. There was nodoubt about this attribution of increased loads ofwater-soluble nitrates to rivers and lakes (Lund,1972). However, it was also shown that the phos-phates applied to the land were rapidly immobilisedin the soil, mainly by clay particles, iron, aluminiumor calcium salts, and so scarcely leached at all bydrainage water (Cooke, 1976). Similar properties ofsoils adjacent to septic tanks provided a similarlyeffective barrier to the transport of phosphorusderived from domestic sewage in rural areas. Bycontrast, there was an increasing application ofsecondary treatment of municipal sewage, the mainpurpose of which was to overcome the high bio-


chemical oxygen demand and associated health risksposed by its organic and bacterial content. Thisprocess fails to remove much of its original phos-phorus content (70–100% is now known to pass theworks : Ka$ llqvist & Berge, 1990) and, moreover,delivers most of it to receiving waters in the dissolvedform most readily available to uptake by aquaticplants (Wilson, 1977). In addition, the mid-centuryexpansion in the use of detergents based on sodiumtri-polyphosphate (STPP, Na

&P$O

"!), which hydro-

lyses to readily available phosphate (Clesceri & Lee,1965), together with its penetration into the newsewage-treatment systems, helped to compound theview that phosphorus enrichment of water coursesemanates primarily from urban settlement.

The many attempts to reverse the water-qualityimpacts of eutrophication by reducing phosphorusloads have provided additional focus on the be-haviour of phosphorus. Whereas there are goodexamples of restoration of individual lakes as aconsequence of sharply diminished point loads (bydiversion or through the application of ‘phosphate-stripping’ tertiary treatment of the final effluent,and assisted by public espousal of allegedly environ-ment-friendly, phosphate-free detergents), the re-sults have not been immediately successful every-where (Sas, 1989). Only modest impacts have beenmade upon the large amounts of phosphorustransported in rivers to the coastal shelf waters,which are now themselves showing symptoms ofeutrophication (Gibson, 1997). This outturn had notbeen anticipated: there is an understandable concernto establish the extent to which internal cyclingaccounts for the apparent insensitivity of someaquatic systems and to separate the impacts of thediffuse, non-point sources of nutrients.

Both problems require a sharper awareness of thechemistry and transformability of the phosphoruscompounds reaching natural water bodies, of theimpact of their speciation on the biological avail-ability of phosphorus to primary-producer biomassand, thus, of its precise role in promoting eutrophi-cation.

(4) Magnitude of phosphorus loads

It is helpful also to have an appreciation of the scaleof phosphorus loadings to some impacted systems. Ina recent survey of 98 major rivers of England andWales, Muscutt & Withers (1996) found that medianconcentrations of dissolved, reactive forms of phos-phorus (see Table 1) exceeded 100 mg P m−$ in 78 ofthem and, in 16 of them, they recorded values in

excess of 1000 mg P m−$. For comparison, concen-trations in many of the world’s greatest lakes rarelyexceed 1–2 mg P m−$, while the OECD thresholdlevel of lake eutrophy is placed at 35 mg P m−$.Closer inspection of the data of Muscutt & Withers(1996) reveals that the highest mean concentrationsare associated with rivers draining the relatively mostpopulous catchments, several of which (the Wear,Trent, Thames and the Bristol Avon) yielded annualaverages equivalent to some 300–600 kg P km−#. Incertain small urban tributaries of the Thames, arealyields exceeded 1000 kg P km−#. All receive majorpoint-source inputs from sewage treatment works,reinforcing the linkage between large loads andsubstantial concentrations of humankind.

In her analysis of data compiled from 32 majorworld rivers (including the Amazon, Congo, StLawrence, Mississippi and Volga), Caraco (1995)found a statistically-robust relation between theexport of dissolved phosphorus (range: 0.3–300 kg P km−# year−") and human population den-sity of the catchment (persons km−#). The slope ofthe regression she fitted suggests an average yield of0.2 kg per individual (ind) per year. When this iscompared with the several estimates of the per capita

physiological generation that are available, however,it is clear that much of the phosphorus output of thehuman population fails to reach the rivers con-cerned. The most widely-accepted estimates areprobably those of Morse, Lester & Perry (1993),who cite 1.6 g P ind−" day−"(i.e. 0.58 kg P ind−"

year−") as a standard output. Historic detergent usemay have accounted for a further 0.7 kg P ind−"

year−", while up to 0.3 kg P ind−" year−" may beattributed to other household wastes in sewage.Moreover, Caraco’s (1995) analysis shows that thehighest area-specific loads emanate from urbanisedcatchments in western Europe (Po, Tiber, Thames,where the rivers discharge dissolved phosphorus atrates equivalent to more than 0.7 kg P ind−" year−").By contrast, rivers draining populous catchmentslacking substantial sewer systems (Hwang, Yangtse,Ganges) transport less than the equivalent of0.1 kg P ind−" year−". Thus, the preliminary evi-dence is that the dissolved phosphorus in rivers mayemanate predominantly from secondary sewagetreatment.

A counterview that non-point sources of phos-phorus, generated primarily in agriculture, mightnevertheless play a significant and increasing role inthe eutrophication of the receiving waters has beengaining credence and has received responsiblesupport. In the USA, losses of phosphorus from land


in cultivation have long been recognised andacknowledged to exceed considerably those fromforested catchments (Hobbie & Likens, 1973;Harper, 1992). Moreover, the rates of P loss havebeen shown to increase as the proportion ofcatchment under forest decreases (Omernik, 1977);both the soluble and particulate loads are affected.Intensification of fertiliser application to agriculturalland in Finland has resulted in significantly accel-erated P-loss rates to water (Rekolainen, 1989). In arecent study of the catchment of Lough Neagh,Northern Ireland, Foy et al. (1995) demonstratedthat areas of established intensive agriculture havebeen exporting increasing loads of dissolved phos-phorus, despite the simultaneous application oftertiary treatment of domestic sewage effluents.Indeed, Lennox et al. (1997) showed that the solublephosphorus not accounted for by the point sourceswas equivalent to some 70–100 mg P m−$, with,interestingly, the higher concentrations matchingthe wetter years. An underlying upward trend ofdissolved phosphorus has been noted in otherNorthern Ireland catchments : expressed as an area-specific loss rate, the rise in soluble phosphorus isequivalent to a year-on-year increase of 10 kg P km−#

year−" (Foy and Withers, 1995). The loss is relativelytrivial in agronomic terms but the concern is stillthat an increasing proportion of bioavailable phos-phorus in water courses is supplied from agriculturalland.

A similar conclusion came from a recent study ofrivers and lakes in the Netherlands (van der Molenet al., 1997). In a notable attempt to apportion thesources of the 105 million kg year−" of phosphorustransported to the Black Sea by the Danube, Brunner& Lampert (1997) attributed almost half (58million kg P year−") to agriculture. Despite beinghome to 70 million people, as little as 23 million kg Pyear−" emanates directly from domestic sewage(approximately 0.3 kg ind−" year−") ; a further 14million kg P year−" comes from industrial processes.Of the load attributed to agriculture, almost half ofthat was attributed to soil erosion; the extent of itsbioavailability has not been established.

This cursory overview confirms the scale ofanthropogenic acceleration of biologically activephosphorus loads to drainage waters. It also revealsthat assumptions concerning their principal prov-enance in secondarily treated urban waste waterneed to be questioned while the validity of pastassumptions about the phosphorus retentivity of soilscontinues to provoke doubt. It is important toquantify the significance and biological activity of

the phosphorus that is potentially leaked to freshwaters by their catchments, not least because it islikely to be much more difficult to regulate than thepoint-sources of the element (Sharpley & Reko-lainen, 1997).

II. CHEMICAL SPECIATION OF PHOSPHORUS

IN NATURAL WATERS

(1) Sources

Phosphorus (atomic weight : 30±974) is the twelfthmost abundant element in the lithosphere (Toy,1973). It occurs in combination with metals (phos-phides), halogens (phosphorus halides, phosphorylhalides), sulphur, nitrogen and, most familiarly,oxygen. In all the biologically active fractions and inall the important minerals from which they areobtained, phosphorus is embodied in the ions oforthophosphoric acid, OP(OH)

$(Emsley, 1980).

Orthophosphoric acid is a weak tribasic acid,colourless and freely soluble in water. The relativeproportions of its various anions (PO

%

$−, HPO%

#−

and H#PO

%

−") vary with pH. The hydrogen radicalsare all replaceable by metals. The orthophosphatesof the alkali metals (with the exception of lithium)are also soluble but those of the alkaline earth andtransition elements – notably, those of calcium,aluminium and iron – are quite insoluble. It is inthese geochemical forms that most of the world’sphosphate resides. The crystalline apatites, par-ticularly fluorapatite [Ca

&(PO

%)$F] and hydroxyl-

apatite [Ca&(PO

%)$(OH)], are of commercial impor-

tance; phosphorite, another commercially exploitedsource of calcium phosphate, is mainly amorphous.The bioavailability of these low-solubility phosphatesin drainage water is normally minimal, subject onlyto gradual weathering and acidic leaching.

The bioavailability of orthophosphate is alsosignificantly affected by the exchange of its ions forhydroxyls on the surfaces of metal oxides and(especially those of aluminium) where they areimmobilised (adsorbed). They are also liable to betightly bound into clay lattices or co-precipitatedwithin, or occluded (absorbed) by, the molecularmatrices of metal oxyhydrides. It is rarely simple todistinguish adsorption and absorption: the term‘sorption’ is used to refer to the association of ionswith solid phases, without precise distinction.

The sorption-desorption reactions between ortho-phosphates and redox-sensitive metals, such as ironand manganese, are especially important in affectingspeciation in aquatic systems. At redox potentials


below 200 mV, the higher-oxidised ion, Fe$+, isreduced to the divalent state (Fe#+). Whereashydrolysis of the trivalent ion leads to the pre-cipitation of the insoluble ferric hydroxide, divalentferrous ions remain in solution. Raising the redoxpotential causes the opposite reaction (Fe#+®eUFe$+), with the iron coming quickly out of solution asa floccular precipitate of Fe(OH)

$, scavenging

dissolved orthophosphate ions as it does so. Affinitiesare such to result in substantial immobilisation ofthese ions, to the extent that they are scarcely anylonger available to biological uptake, pendinganother lowering of redox with re-reduction of theiron to the ferrous state (Fe#+). The phosphorus thus‘released’ back into solution is, potentially, oncemore bioavailable (Golterman, Bakels & Jakobs-Mo$ glin, 1969).

Redox-mediated changes in phosphate solubilityat the bottoms of lakes were described over sixtyyears ago (Einsele, 1936; Mortimer, 1941, 1942).Since then, many of the beliefs about the impacts ofphosphorus enrichment on aquatic ecosystems havecontinued to be dominated by concerns about itsspeciation and the availability of phosphorus inaquatic systems. Solubility transformations affectedby pH and by the behaviour of other elements(Lijklema, 1977) may be just as important to thebiological responses in aquatic environments. Thecomplexity of these transformations and the quantifi-cation of their effects continues to challenge theunderstanding of phosphorus bioavailability.

(2) Phosphorus fractionation and biologicalavailability

The biological importance of phosphorus, as acomponent of nucleic acids and for its pivotal role inintracellular molecular synthesis and transport,ensures that the biotic transformations of the elementconstitute a vital linkage in its geochemical cycling.These transformations involve exclusively the ortho-phosphate radical. Phosphates may be returned tothe environment in solution, as in animal metabolitessuch as mammalian urine, or as the relativelyinsoluble calcium salt that comprises some 60% ofvertebrate bone. Phosphorus in plant and animaltissues is linked to carbon through C-P or C-O-Pester bonds (Engstrom & Wright, 1984; Baldwin et

al., 1995). Phosphate liberated from decomposingdead tissue may be sorbed or complexed with metaloxides and clay minerals and, thus, subject totransport and recycling in the same ways asphosphate from other sources (see later).

The relative insolubility of certain metal phos-phates and the variety of sorption reactions leadingto their immobilisation mean that free phosphate isoften a scarce resource in functional ecosystems.Biological availability (or bioavailability) of phos-phorus refers, in essence, to those fractions of thetotal mass of phosphorus present in a system that arereadily assimilable by organisms, or are made moreassimilable through the activities of the organismsthemselves (for instance, through the production ofphosphatases), and that portion of the phosphoruswhich has been assimilated already and is intra-cellular. In aquatic systems, bioavailable phosphorusis introduced primarily as a consequence of aqueousleaching from the hydrological catchment. Thequantities of usable phosphorus available are relatedfirstly to the geochemistry of the catchment and themineral deposits present, then to rates of theirerosion, which will be influenced by geomorphology,climate and hydrology. The extent of sorption ofliberated phosphate by catchment soils and of itsassimilation into terrestrial biomass will furtherinfluence the amounts transferred to water. Thephosphorus reaching streams, rivers and lakes can beorganic or inorganic, and in solution or as particu-lates of various sizes. Once in water, the furtherbiological partitioning of the phosphorus and theextent of its immobilisation in the bottom sedimentswill be greatly influenced by the distribution amongthe fractions in the imported influx, to its within-water chemical transformations and the extent ofinactivation through sediment sorption. Quite nat-urally, the concentrations of the biologically relevantforms of phosphorus among fresh waters range overfive orders of magnitude (Reynolds, 1984).

The forms of phosphorus present in aquaticenvironments and in the drainage waters supplyingthem may be dissolved, colloidal or in particles. Fewof the particulate forms are readily bioavailable. Thecomplexities of phosphorus geochemistry make itdifficult to judge the functional impacts and man-agement implications of the totality of phosphoruspresent. However, the adoption of analyticallydefined fractions has proved helpful to estimatingtheir bioavailability, according to their chemicalreactivity. Sequenced, serial analyses are used toseparate the fractions in which the phosphorusresides, removing them stage by increasingly ag-gressive stage. The first such fractionations wereapplied to soils (Deans, 1938; Chang & Jackson,1957) but they have been progressively developedfor lakes (see, among others, Williams et al., 1971;Williams, Jacquet & Thomas, 1976; Hieltjes &


Lijklema, 1980; Psenner et al., 1988; Baldwin, 1996;Davies, 1997) or sea water (Ruttenberg, 1992).Approximations of the chemical compositions andbioavailabilities of the main ‘operationally-definedfractions ’ are summarised in Table 1.

The ‘readily bioavailable ’ fractions in water arethose in solution and measurable directly in waterpassed through fine (0.5 µm) filters. They includethe fully dissociated orthophosphate ions (H

#PO

%

−,HPO

%

#−, PO%

$−) that are registered by the standardanalytical determination of phosphate phosphorusby the molybdenum-blue method of Murphy &Riley (1962). A yellow complex between ortho-phosphate and ammonium molybdate is formed andreduced with ascorbic acid to give a blue solution.This method is also sensitive to the presence ofphosphate held in colloids so, in many instances, thismolybdate-reactive fraction exceeds the mass ofphosphorus in true solution (Rigler, 1968). It ispossible to distinguish the ‘dissolved phosphorus’(DP) fraction by microfiltration, though this israrely applied. The ‘molybdate-reactive phos-phorus ’ (MRP) or ‘ soluble, reactive phosphorus’(SRP) fraction measured in raw or coarse-filteredwater samples is usually wholly removable bybiological uptake, at least to the limits of itsanalytical detection: MRP is to be considered‘readily bioavailable ’.

Natural MRP concentrations are frequently belowthe detection limits of the earliest analytical methods(Lund, 1965). By the time that the sensitivity of themolybdenum-blue method had been increased ad-equately (by extraction and concentration of themolybdenum blue in an organic solvent : Procter &Hood, 1954), it had been realised that the bio-available fraction exceeded, sometimes considerably,that which was measurable as MRP. Estimation ofthe biologically relevant phosphorus pool needed toinclude the organismically bound fraction. Thus,methods involving the determination of orthophos-phate in solution after oxidation of the organicfraction (Mackereth, 1963) were widely adopted.The perchloric-acid digestion of samples to yieldmeasurements of their ‘ total phosphorus’ (TP)content has become the standard in establishing thefertility and potential quality of water ever since.

It is important to emphasise (mainly because it isso frequently overlooked) that TP often includes alarge proportion that is not immediately or necess-arily ever bioavailable. Intrabiotic components arereadily available. Of the extrabiotic non-dissolvedTP, some could become available under definedenvironmental conditions (see below); others (such

as the phosphates of alkaline-earth metals) areeffectively unavailable. The bioavailability of this‘particulate phosphorus’ (PP) fraction (i.e. thatwhich is removable from water samples by 0.5 µmfiltration) is estimable retrospectively, after the seriesof steps applied to extract reactive orthophosphatefrom the particulate material has been applied. Theparticulate fraction that is already organismic, plusthat available in solution, may then be collectivelydistinguished as ‘biologically available phosphorus’(BAP).

The ‘conditionally-bioavailable particulate phos-phorus (Table 1) includes phosphate so looselybound that it is readily desorbed and dissolved intothe water filling any interstices of the particulatematerial and which, once separated from the solidmatter, is measurable as DP or MRP. The conditionof its availability is simply that the water is releasedinto, or exchanged with, the water exterior to theparticles. It is identified in Table 1 as ‘ interstitialmolybdate-reactive phosphorus’ (IMRP). Con-ditionally-available phosphorus also includes sorbedorthophosphate ions that are exchangeable withchloride ions. This, the most readily-available of theparticle-bound fractions, is sensitive to the ionicstrength of the bathing water. It is removedanalytically with 1 M ammonium chloride solutionand referred to as ‘ammonium chloride-extractablephosphorus’ (NH

%Cl-P).

Next are the redox-sensitive iron- and manganese-bound fractions. Adequate reduction of ferric hy-droxide in bicarbonate-buffered medium is attainedwith a solution of sodium citrate and sodiumdithionite. The ‘citrate-dithionite-extractable phos-phorus ’ (Na

#S#O

%®P) represents the phosphorus

sorbed or occluded by iron which is normallyreleased into solution under conditions of loweredoxidative capacity. The pH-sensitive aluminium-bound phosphorus is exchangeable in molar sodiumhydroxide, the ‘ sodium hydroxide-reactive phos-phorus ’ fraction (NaOH-rP).

The remaining particulate phosphorus is relativelystable and scarcely bioavailable without some pro-tracted diagenesis of the material. Davies (1997)referred to these recalcitrant fractions generically asbeing ‘ sodium hydroxide non-reactive ’ (NaOH-nrP): although chemically unexplicit, it is a helpfulclassificatory term. Part of this is apatitic or is co-precipitated with calcium carbonate and is releasedonly by treatment in strong acid to dissolve thecarbonates and oxides. This ‘hydrochloric acid-extractable phosphorus’ (HCl-rP) fraction is de-termined by dissolution in 1-M acid. The remainder,


the refractory, organically bound ‘residual P’ inTable 1, is the least well understood. Some of themain potential pathways of bioavailable phosphorusinvolve biogenic metabolites and biomass residues.Some of these, for instance, the phospholipids,nucleic acids and nucleotides, sugar phosphates andinositols, are quite labile and come out with thereadily hydrolysable and exchangeable fractions(Morgan, 1997) but more resistant forms arecomplexed with humic acids (Sharpley & Reko-lainen, 1997). Humic acids usually represent thelargest fraction of organic matter in dissolved waterdraining from natural catchments (Wetzel, 1995;Stumm & Morgan, 1996) where they act aspolyelectrolytes for anionic functional groups. Lig-and exchange of carboxyl and phenolic groups onthe humic molecules for surface hydroxyls is thoughtto occur. Resultant colloids can immobilise organicP molecules (such as nucleic acid: Crechio &Stotzky, 1997) as well as those sorbed to ironhydroxides (Francko, 1990; Jones, Shaw & de Haan,1993). The amount of humic material present maythus regulate the formation of iron floc and,consequentially, the conditional bioavailability ofphosphorus (Shaw, 1994). The sensitivity of humicacids to ultra-violet wavelengths may be importantto this regulation.

(3) Relative distribution of phosphorusfractions in natural waters

Few activity constants are known to govern thetransitions among the various phosphorus fractionsin natural waters (Lean, 1973): absolutely andrelatively, the composition of natural waters withrespect to the bioavailability of phosphorus isextremely varied. Distinctions among water typesare most usually made on the basis of their totalphosphorus (TP) contents, which range pre-dominantly between 10−) and 10−% M (0.3 mg to3 g P m−$).

Few detailed inventories of the phosphorus frac-tions transported through lake and river catchmentsare available in the literature. Mostly, the necessaryserial analyses are so tedious and, ultimately, soimprecise, that many workers have been content todistinguish from the total phosphorus (TP) value, anSRPfraction, an NaOH-rP fraction and, perhaps, anHCl-rP fraction (Davies, 1997). General patternsmay be recognised. Well-developed terrestrial veg-etation tends to accumulate readily-bioavailablephosphorus, exports to water being confined largelyto humic-bound leachates (Taylor, Edwards &

Simpson, 1971; Hobbie & Likens, 1973). The moreis the catchment subject to physical erosion, thegreater is the transport of particulate phosphorus,although much of this must remain biologicallyunavailable (Omernik, 1977).

The challenge remains to confirm and to quantifythe extent of the biological exploitability of thevarious fractions in drainage water. Assumptionsabout the soluble, MRP forms being universallyavailable to algae, bacteria and many kinds of waterplant are upheld by the results of numerouslaboratory experiments in which MRP is removed,within hours, from solution by P-deficient algae(Lean, 1973; Button, 1985; Davies, 1997). Manyplanktonic algae show high affinity for phosphorus(Fisher & Lean, 1992) down to very low thresholdconcentrations, in the nanomolar range (Falkner,Falkner & Schwab, 1989). Similar experiments havebeen devised to measure the use of the less reactivephosphorus fractions, in which algal uptake orgrowth becomes a means of assaying the bio-availability of the phosphorus provided. A thoroughreview of this topic (Ekholm, 1998) concludes thatTP is a poor indicator of algal-availability, whereasMRP was nearly always a measure of the minimumavailable. The phosphorus afforded by the lessreactive P forms depends upon the nature of theconditional availability, and the circumstances gov-erning the direction of the sorption}desorptionreactions.

In general, the bioavailabilities of the phosphorusfractions match their chemical reactivities quite well.Matters are less clear in the case of the organic-bound P fractions – clearly, there are many kinds oforganic P and there are many organisms of un-measured capabilities, but the facility for secretingalkaline phosphatases with the specific function ofcleaving orthophosphate groups from certain typesof organic compound is well established (see Cem-bella, Antia & Harrison, 1984). This provides anexception to the general correspondence betweenchemical reactivity and bioavailability.

III. THE ASSIMILATORY VALUE OF AQUATIC

PHOSPHORUS FRACTIONS

(1) The Vollenweider-OECD model

There is no more powerful or elegant representationof the apparent consistency of system responses toenrichment than the Vollenweider-OECD phos-phorus-chlorophyll regression, commonly referred toas ‘ the Vollenweider model ’. The final form of the


10

1

1 10 100

An

nu

al m

ean

ch

loro

ph

yll a

co

nce

ntr

atio

n (

mg

m–3

)

[L(P)/qs]/[1+(z/qs)0.5]

Annual index of phosphorus availability (mg m–3)

99%

r = 0.868

Fig. 1. The Vollenweider–Kerekes regression relatingobserved annual mean chlorophyll-a (chl)concentrationin northern-hemisphere lakes to an annual index ofphosphorus availability [P

y], based upon phosphorus load

[L(P)], mean depth [z] and qs

is the reciprocal of thehydraulic retention time, t

w. The equation of the fitted

regression is : log [chl]¯ 0.91 log [Py]®0.435. Redrawn

from Vollenweider & Kerekes (1980).

regression (Vollenweider & Kerekes, 1980) is repro-duced here (Fig. 1) in order to make a number ofpoints about its strengths and to highlight some ofthe ways in which it is misapplied in the managementof eutrophication. The regression provides an im-portant statement of the fact that, within the limitsof common experience of a wide range of temperatewater bodies, phosphorus is scarce to the extent thatthe more of it that is available, the greater is theaverage biomass (for which chlorophyll content is aconvenient analogue) of planktonic algae supportedBased, as it is, on a large set of data (albeitpredominantly from temperate northern-hemispheresystems), the regression certainly attests to thegeneralised, long-range behaviour of a large numberof the sites considered.

The problems we allege are not the fault of themodel ; they relate to its simplistic interpretation.First, there is an implicit presumption that onlyplanktonic algae respond to phosphorus enrichment.True, in larger lakes, well away from the marginsand the bottom, the phytoplankton is the onlysubstantial producer biomass. Currently, there is nowell-tested regression which relates the biomass ofemergent and submerged herbs to nutrient avail-ability. Neither is account often taken of the fact thatthe plant biomass nourishes higher trophic levels ofconsumers which, conversely, erode the base of plantbiomass. Nevertheless, several authors have attrib-

uted lower-than-predicted levels of average plantbiomass to persistent herbivory (Rosenzweig, 1971;Canfield & Bachmann, 1981; Makarewicz &Bertram, 1991). By analogy, consumption of zoo-plankton by planktivores could compensate in theopposite direction. In fact, positive relationshipshave been demonstrated between long-term aver-ages of zooplankton biomass and phytoplanktonchlorophyll (McCauley & Kalff, 1981), and betweenthe biomass of planktivorous fish and zooplankton(Mills & Schiavone, 1982). Moreover, the averagerelation of chlorophyll concentration to total phos-phorus loading is erratic in lakes where there arerelative deficiencies in nitrogen (van Donk et al.,1993) or there is light limitation of biomass, as aresult of mixing depth, of high turbidity or, at thehigh algal concentrations associated with highavailabilities, of self-shading (Reynolds, 1992). Sev-eral authors (Prairie, Duarte & Kalff, 1989; Foy &Withers, 1995) are agreed that the correlationbetween chlorophyll concentration and phosphorusavailability breaks down at loadings sufficient tomaintain total phosphorus concentrations (TP)exceeding 100 mg P m−$. Indeed, a ‘ sigmoid re-lationship’ between chlorophyll and phosphorus hasbeen proposed (McCauley, Downing & Watson,1989; Prairie et al., 1989; Watson, McCauley &Downing, 1992): this alludes to a steep directrelationship above one threshold and then a levellingoff at a higher one. This is perfectly reasonable butthe relationship is sigmoid only when plotted onnormal scales : it clearly cannot hold for the log}logformat used by Vollenweider & Kerekes (Fig. 1).

Taking into account the above discussion aboutthe Vollenweider}OECD regression, its value as aprecise predictive management model is quite un-promising (Vollenweider, 1976). Yet it is clear thatmanagers have been prepared to invest in controlsbased on a target chlorophyll level that is an annualaverage yielded by a log-log regression, fitted to alimited selection of well-studied lakes, and whichitself is subject to a ³50% error. Besides, at highbiomasses, phosphorus and chlorophyll may both beintracellular dependents of algal abundance.

(2) Cell stoichiometry

A quite different approach is needed to determinethe supportive capacity of the phosphorus load onlakes and rivers. The most desirable outcome is aload-specific biomass yield. This is less easy toestimate than it is to calculate an instantaneousbiomass carrying capacity. The latter is needed to


develop the former. The essential logic is based onthe supposition that when all known requirements ofthe cells of planktonic algae (and, indeed, the cells ofall living organisms) are optimally supplied, theyhave similar and predictable elemental constitutions,approximately in the molecular ratio recognised byRedfield (1958). Ignoring liquid water and bio-mineralised structures, biomass comprises 106 atomsof carbon and 16 of nitrogen for every one ofphosphorus. In terms of mass, the ratio approximatesto 41:7:1. Then every gram of bioavailable phos-phorus in water has the capacity to support 41 g ofplant carbon. Taking Reynolds’ (1984) regressionequations, it may be shown that the ash-free drysolids of non-vacuolate planktonic algae account forsome 41–47% of the live mass. Of this, 51–56% iselemental carbon. Against a typical composition of0.21–0.26 g C (g live mass)−", the same material com-prises 0.03–0.04 g N and 0.005–0.006 g P (g C)−"

(note, phosphorus represents approximately 1% ofthe ash-free dry mass).

These numbers simplify the calculation of phos-phorus-carrying capacity. Neither biovolume nororganismic carbon are easily measured, however, sotheory is not readily verified. As an index of activeprimary-producer biomass, however, chlorophyllpigments (especially chlorophyll a) have manyanalytical attractions. Chlorophyll a is a functionalcomponent of all active photoautotrophs, there is awell-tested method for its extraction and measure-ment and there are far more data on algalchlorophyll than there are about even the organismsin the original samples whence the chlorophyll isextracted. Despite an acknowledged wide variabilityin how much chlorophyll there is in an algal cell(0.5–5% of ash-free dry mass), it has been foundthat nutrient-replete cells kept under moderateintensities of artificial light generally tend to achlorophyll content of approximately 1% of ash-freedry mass. The approximate 1:1 equivalence withphosphorus is fortuitously very useful to empiricalapproximations. The actual chlorophyll contents ofthe algae in the dataset analysed by Reynolds (1984)ranged from 0.9% to 3.9% of dry weight (1.6–7.7%of cell carbon; cells had carbon-to-chlorophyll ratiosof 13:1 to 63:1).

Against reported phosphorus contents rangingbetween 0.7 and 3.0%, there is scope for con-siderable variation about the 1:1 parity (Reynolds,1992). Clearly, if the theoretical stoichiometry is tobe of practical use, some further verification isneeded to reduce the uncertainty. By fitting data onchlorophyll yields of natural populations to the

1000

100

10

1

1 10 100

Max

imu

m c

hlo

rop

hyl

l co

nce

ntr

atio

n (

mg

m–3

)

Available phosphorus (mg m–3)1000

Fig. 2. Regression describing maximum chlorophyll-aconcentration ([chl]

max) measured in each of a selection of

lakes in north-west England as a function of theconcentration of biologically available phosphorus (BAP).The equation of the fitted regression is log [chl]

max¯

0.585 log [BAP]0.801. Redrawn from Reynolds (1992).

consumption of soluble reactive phosphorus duringdevelopment, Reynolds (1978) did establish a stat-istically sound regression which has succeeded inaccommodating all the data subsequently tested(Reynolds, 1992). The equation of the line, shown inFig. 2, is instructive for several reasons. One is thatit is sublinear : a phosphorus availability equivalentto 10 mg P m−$ is predicted to have the potential tosupport phytoplankton containing a maximum of24 mg chlorophyll a m−$ ; for 100 mg P m−$, themaximum yield is 94 mg chlorophyll a m−$ ; for1000 mg P m−$, it is 359 mg chlorophyll a m−$. Inother words, the efficiency of biomass assemblydeclines with the relative abundance of the resource,or when the supply of other resources becomecritical. This is adequately realistic (Gibson, 1997).

The second point is that the relationship is scarcelyaffected by phosphorus other than that in the‘readily available ’ fraction; i.e. the biomass responseis to BAP and not to TP. Once again, TP is seen tobe a poor predictor of biological availability whileMRP is a rather useful measure of the unusedphosphorus-determined carrying capacity of a givenwater.

Stoichiometric elemental ratios have been usedwidely to identify ‘ limiting factors ’ in systems.However, a more fruitful derivation is the calculationof the carbon-assembly capacity provided by theinstantaneous availability of each nutrient. The


carbon supply itself and the capacity for its re-oxidation can be similarly treated, as can thecarrying capacity of the underwater light field(Reynolds, 1997). Besides identifying the likely limitsto system metabolism, this process also reveals thosevariables to which the biomass is presently in-sensitive. This particular approach has not beenmuch tried but it offers a measurement of the extentto which system outputs are continuous functions ofa single resource (phosphorus availability, say) orare sensitive to variable capacity limitations by otherconstraints (carbon delivery, light, other nutrients).In these ways, we may recognise that a BAPconcentration equivalent to only 3 g P m−# is per-fectly capable of supporting phytoplankton up to anenergy-limited capacity of 50 g C m−#, while aphosphorus flux equivalent to 7 g P m−# year−",supported by a nitrogen supply of 50 g N m−# year−"

is sufficient to leave the system highly dependent onthe flux of carbon dioxide.

The ability of rooted aquatic macrophytes to storeexcess photosynthate and to accumulate carbon instructural components (xylem, sclerenchyma) of themechanical ‘necromass ’ compounds their ability todraw directly on sedimentary sources of phosphorus.Moreover, aerial structures, linked via aerenchy-matous tissue to the sub-surface organs, afford toemergent macrophytes greater energy- and carbon-determined thresholds to supportable biomass ca-pacities : we should always expect them eventually tobe better equipped than algae to dominate shallow-watered habitats, other factors notwithstanding.

(3) Phosphorus uptake

The nutrient-uptake capacity of biological systemshas been an important focus of physiological ecologyfor over 40 years. As already noted, nutrient-starvedalgal cells have high affinity for bioavailable phos-phorus and they take it up rapidly (Button, 1985). Ithas been for long appreciated that uptake conformsto Michaelis–Menten-type kinetics whose depen-dence upon affinity and resource concentration aredescribed by Monod models (Dugdale, 1967). Droop(1973, 1974) sought to improve the relationships byincluding the variable internal nutrient content ofcells. His variable-stores model continues to berecognised as the most useful representation ofbiogenically mediated nutrient exchanges. Bur-master’s (1979) important investigation demon-strated a steady-state equivalence among the rates ofnutrient uptake and consumption in growing algae.The various approaches may be combined in a single

relationship between growth (as the rate of specificincrease, r«) and the nutrient available :

r«¯ r«max

(q®q!)}K

U(q®q

!) (1)

This clever formulation supposes that the growingcells have resource in hand (q being the internalstore, or cell quota, and q

!being the minimum quota

on which the cell can survive) so long as this is madegood by uptake, and which is itself half-saturated ata concentration, K

U; r«

maxis the maximum growth

rate and, like KU, is determined by experiment.

The perspective provided by modern molecularapproaches to nutrient acquisition helps us tounderstand how the model works and how a stillsimpler model might be applied. The movement ofinorganic materials into and within the cells iseffected principally through a series of highly specificmembrane transport systems. They conform to ageneral structural arrangement that is found in thematerial- and information-conducting systems of allliving cells. They are generally powered by iongradients and proton motive forces and they arefuelled by ATP-phosphorylation (Simon, 1995). Inthe case of nutrient uptake, for example, the circuitsinvolve sequences of protein-protein interactionswhich convert the binding of a ligand at a peripheralreceptor to an excitation of the transfer response.The receptor region of one of these complexes isperiplasmic and constitutes a ligand-specific protein.Reaction with the target molecule stimulates amolecular transformation, which, in turn, becomesthe excitant substrate to the proteins of the trans-membrane region. Sequential analogous reactionsconduct the nutrient molecules to the intracellularsites of consumption along a redox-gradient ‘chan-nel ’. It is clear that the channel can be underused (asduring a shortage of target nutrient molecules) orsaturated (when nutrient is supplied in excess, andthe transport system is restricted to accepting targetmolecules only as fast as they can be consumed). Atthe same time, the level of activity of the transportmechanism is signalled to the genes controllingassembly and regulating cell division, and a per-sistent shortage will necessitate the implementationof the cell’s self-conserving reactions to nutrientstarvation (Mann, 1995).

These findings vindicate the Droop–Burmastermodel in almost all respects. However, the naturaldeficiency of phosphorus in aquatic environments ismet by the adaptation of planktonic algae to be veryefficient in pulling in their phosphorus requirements.All planktonic species so far examined are able totake up sufficient phosphorus to saturate the phos-


phorus demands of the fastest attainable growthrates from external concentrations containing SRPat less than 10−( mol l−" (3 mg SRP m−$ : Reynolds,1997).

Thus, the Monod uptake curve describes whatamounts to a massive over-capacity for algae to beable to take up available phosphorus. This is animportant adaptive asset in phosphorus-deficientenvironments but its corollary is that the immediategrowth prospects for algae, maintaining Redfieldstoichiometry, are scarcely impaired at SRP con-centrations exceeding 10−( mol l−". For the purposeof practical deductions, three further corollaries maybe derived. One is that if phosphorus is present in thewater at concentrations much in excess of 3 mgSRP m−$, it cannot be considered to be growth-limiting. Second, growth can proceed independentlyof the external phosphorus concentration until theSRP is reduced almost to the limits of its analyticaldetection. Third, for so long as SRP exceeds thislevel, the phosphorus-determined supportive ca-pacity of the medium is under-filled.

Many species tolerate chronically P-deficientconditions by producing alkaline phosphataseswhich sequester phosphate from certain kinds oforganic compounds (Cembella et al., 1984). Mixo-trophy (supplementing autotrophy by bacterivory)provides a valuable alternative source of nutrients tocertain kinds of algae (see Riemann et al., 1995, fora review). Such nutritional modes are clearlyimportant to individual organisms in locations whereSRP is frequently or chronically in short supply but,for the same reason, their contribution to the adverseeffects of enrichment by excesses of anthropo-genically generated SRP is negligible.

(4) Estimating the bioavailable fraction

The phosphorus available to support algal growth isnot confined to the analytically determined SRPfraction but largely comprises that proportion of theTP that is present in biomass (BP). The idealinstantaneous measure of P bioavailability is BAP¯SRPBP. As already recognised, algal cells arecapable of rapidly removing SRP to store in BP:other things being equal, SRP decreases but TP isconstant. TP should never be taken as a measure ofBAP, especially in oligotrophic to mesotrophicsystems where a significant, if not the largest, fractionof the TP will be present in the scarcely or onlyconditionally available clay- and metal-sorbedfractions. Nevertheless, exhaustion of SRP to ana-lytically undetectable levels confirms that all BAP

has entered the BP fraction and a growth-limitingcondition has been reached.

However, the sum (SRPBP) is difficult todetermine directly. With the benefit of time-sequen-ced samples, biomass (analogised to chlorophyll-aconcentration) may be regressed against TP to solvefor the intercept corresponding to the non-availablePP base. Alternatively, the chlorophyll can beapproximated as a direct stoichiometric correlativeof BP (provisionally, at least, from the 1:1 re-lationship, discussed above), and compounded bysummation with the residual SRP as a minimalestimate of the BAP. Again, it is obvious that SRPexternal to cells represents excess capacity at thetime of its analysis and that all other chemicallybound forms of non-organismic P are of littleimmediate biological consequence while measurableconcentrations of SRP persist.

IV. TRANSFORMATIONS AND

TRANSLOCATIONS OF PHOSPHORUS IN

AQUATIC SYSTEMS

(1) Proximal fates of phosphorus loads inaquatic systems

The above deductions are complicated by the manychemical and physiological transformations thatphosphorus loads undergo within aquatic environ-ments. Unlike silicon and many other biomineralswhose cycling occurs at geochemical scales, phos-phorus is transformed and recycled on relativelyvery short time scales. This section seeks to sum-marise the main aquatic pathways.

The phosphorus entering discrete water courseshas several proximal sources and it is variablypartitioned among the several possible fractions.Most of the load is water-borne, either in solution(sensu SRP) or in particulate suspension (PP in Table1). Dry deposition of wind-blown dust and soil maybe locally significant but, because the phosphorus-containing particles are as likely to be eluted fromthe atmosphere by precipitation, it is not necessaryto distinguish the dry flux from phosphorus leachedby falling raindrops. TP in rain varies with theprovenance of the incoming air masses and the factorof rainfall dilution (see Allen et al., 1968). Ahl (1988)estimated the annual elution in the remoter regionsof the northern hemisphere to be equivalent to some5–6 mg TP m−# year−". With a mean global pre-cipitation of 640 mm year−" (calculation of Rey-nolds, 1997), the nominal TP concentration ofsurface waters should approximate 7–9 mg TP m−$.


If allowance is made for evaporation of some of thewater, the expected concentration might be at leastdouble that. When it is considered that TP concen-trations in rain water collected in more populous ormore agricultural areas are up to an order ofmagnitude greater again (20–130 mg TP m−$ ), withmatching annual areal deposition rates (5–80 mgTP m−# year−" : Allen et al., 1968; Williams, 1976;Holtan, Kamp-Nielsen, L. & Stuanes, 1988; Gibson& Wu, 1995), it becomes possible to anticipateunmodified aqueous concentrations perhaps up to250 mg TP m−$. Such levels are simply not achieved,either in natural lakes (OECD, 1982) or even inmany unpolluted rivers (Caraco, 1995).

Part of the explanation for this is that much of theTP arrives as scarcely-available HCl-rP in par-ticulate fractions which, being lost steadily fromsuspension, never reach the steady state anticipated.However, in many large, unproductive lake waters,where hydraulic loads are influenced by directatmospheric exchanges, the reconstructed BAP frac-tion rarely exceeds 5 mg P m−$ (Reynolds et al.,2000). Elsewhere, the chemical composition of oligo-trophic inland waters (inferred BAP! 5 mg P m−$)is, to a great extent, modified by contact with theterrestrial surfaces over or through which a relativelylarge proportion has first flowed. It is apparent,possibly counterintuitively, that participation interrestrial exchanges and ecosystem function canlead to a marked diminution in the BAP contentof terrestrial drainage. In their celebrated study ofthe Hubbard Brook, Hobbie & Likens (1973) showedthat the P loss in streamflow from the densely-forested catchment was equivalent to an export ofapproximately 2 kg TP km−# year−", against anannual rainfall contribution of close to 11 kgTP km−#

and an annual leaf-fall of 190 kg TP km−#. Ahl(1975) also found the mean yield of phosphorus fromSwedish boreal forest catchments to amount to some3–9 kg TP km−# year−". Phosphorus ‘ leakage’ fromthe vegetation is tiny compared to the turnover :forests accumulate and recycle their phosphorusinputs.

While biomass accumulation in high forest wouldappear to explain the terrestrial retention of phos-phorus, the soil and, where appropriate, the subsoilalso tie up phosphorus through sorption onto claysand metal oxides. Forested catchments are un-generous in the P they export to aquatic ecosystems:the natural condition of these water bodies, certainlythose in the temperate latitudes, is properly judgedto be phosphorus-deficient. Yet it is equally apparentthat the yield of TP from catchments may increase

substantially as a consequence of forest clearanceand of the imposition of agriculture (Dillon &Kirchner, 1974; Rekolainen, 1989). Exports mayincrease by an order of magnitude (to approximately100 kg TP km−# year−"), although the fraction thatis augmented is predominantly particle-bound phos-phorus, contributed by eroding soil. Even theorganic components and biogenic residues shed fromcatchments after forest clearance remain scarcelyavailable to aquatic producers.

(2) Processing the bioavailable fractions

The biomass-supportive capacity of the recipientaquatic ecosystem thus depends primarily upon the(usually) small proportion of SRP delivered from thesky or allowed to escape from the catchment. Theaggregate annual load of soluble phosphorus to therecipient water body, Σ

y(SRP), is the sum of the

annual direct deposits, (SRP)dy

, and the annualload from the inflow, Q

y(SRP)

Iy. Thus,

Σy(SRP)¯ a(SRP)

dyQ

y(SRP)

Iy(2)

where a is the area of the recipient water body andQ

yis the annual inflow volume. The area-specific

load, Ly, is equivalent to [Σ

y(SRP)}a.]. The annual

volume discharged from the catchment, Qy, is the

product of the catchment area (A) and the annualprecipitation (P)

y, net of evaporation loss (E)

y:

Qy¯A(P®E)

y(3)

Then, the soluble phosphorus delivered from thecatchment to the water body is approximately :

Qy(SRP)

Iy¯A(P®E)

y¬(SRP)

s(4)

so,

Qy(SRP)

Iy¯A(SRP)

sy(5)

where (SRP)s

is the area-specific soluble fractionand (SRP)

syis the area-specific soluble fraction

actually shed from the catchment in the year.Once delivered to the receiving water, the soluble

load, Σy

(SRP), may remain in solution, or bescavenged by hydroxyl-rich particulates, althoughthe Vollenweider model suggests that, in mostnatural, phosphorus-deficient waters, much is sooneror later assimilated into the biomass of planktonicalgae and bacteria (as BP). The subsequent fate ofthis phosphorus is then dependent on the fate of thebiomass. In basins subject to significant hydraulicflushing, plankton and particulates are washeddownstream, conceivably to the sea. With severeflushing, SRP may be ‘ lost ’ from the water body


without the intervention of chemical or bioticuptake: the greater is the relative rate of dis-placement, the greater is the rate of P loss. However,this phosphorus remains potentially supportive ofbiomass elsewhere.

In more retentive lakes (displacement times of" 10 days), planktonic BP is subject to three main‘sinks ’. That which is not washed from the lakeintact, to become, by analogy, part of the particulateload downstream, is either consumed by hetero-trophs (usually grazing crustaceans, rotifers orprotists of the zooplankton) or eventually settles tothe bottom sediment. Settlement is an inevitableconsequence of the fact that cells of freshwaterphytoplankton are typically heavier than water.Their sinking velocities are orders of magnitudesmaller than the turbulent velocities typical of wind-mixed surface layers, while storm events and waveaction at the margins ensure frequent random-isations of all entrained material. It is thus unusualto find large differences of cell concentration at-tributable to algal sinking, except in the presence ofpronounced density gradients or significant shearboundaries, such as occur adjacent to solid surfaces.Settling phytoplankton can escape across the lowerboundary of the mixed layer, however: loss fromrandomised distribution complies with first-orderdynamics, analogous to dilution, and at exponentialrates predictable from the still-water settling velocityand the mixed-layer depth (Reynolds, 1984). Be-yond the lower boundary, there is no certain return,pending an increase in the extent of the mixing shearor a change in the sinking behaviour of the settlingalga.

The organic carbon fixed by autotrophic phyto-plankton is exploited by heterotrophs in a number ofways. These include: uptake (mainly by bacteria) ofglycolate and other excess photosynthetic inter-mediates released into the water by the algae;parasitic exploitation (mainly by chytrid fungi) ofcell cytoplasm; and consumption, in part or whole,as the food of planktonic herbivores. So far as thetranslocation of phytoplankton BP is concerned,only the latter is consistently important. However,its importance relative to direct sedimentation isvery variable, being influenced by the depth ofwater, the size-range of the dominant species ofphytoplankton and its concentration, water tem-perature and flushing. The partitioning of phyto-plankton among its possible fates varies conspicu-ously among lakes, between seasons and with thetype and abundance of zooplankton (Reynolds,1984). The extent to which herbivores are consumed

by planktivores, including fish, also influences thetransfer of BP from primary to secondary com-partments and, hence, the source and quantities ofexcreta. Planktonic copepods and most fish passphosphorus-containing faecal pellets which, subjectto microbial decomposition, fall to the lake sediment.Planktonic Cladocera produce less discrete digesta,whence leachable nutrients are presumed to be moreimmediately recyclable to the open water.

Ultimately, however, live algal cells, their bac-terised remains and the excreta, corpses and cadaversof their consumers, constitute the autochthonoussedimentary flux. Over a year or beyond, thetotal quantity of biogenic phosphorus sedimented[Σ

y(BP)

sed] will represent an estimable proportion

of the original bioavailable load, [Σy(SRP)], less

that part washed out in solution or in the biomass oforganisms, and that of BP exported in migratory fishor emergent insects. The primary influences on thebiogenic transformations of P are related to hy-draulic residence time and bathymetry. In terms ofthe annual total phosphorus budget, the total load ofall sources, [Σ

y(TP)

Iy], should be balanced by the

total deposition [Σy(TP)

sed], less an export which

should be close to the difference [Σy(SRP)

Iy]®

[Σy(BP)

sed].

Few data are available with which to verify this.As a hypothesis, it holds for the deeper lakes whosephosphorus relationships were reviewed by Sas(1989), for the nutrient budget for Windermere(Reynolds, 1999) and the behaviour of limneticenclosures (Reynolds, 1996). In general, it is reason-able to view the lake as a substantial trap for itsphosphorus load, much of which seems likely to bedestined to become incorporated in the permanentdeep sediments.

(3) The influence of macrophytes

The above deduction does not apply everywhere orat all times. Before considering the exceptions,however, it is important to clarify the role ofmacrophytic primary producers. Plants with dif-ferentiated tissues, not exclusively rooted ones orones with thick rhizomes, are exempt from thenecessity to shed excess carbohydrate : many areequipped to retain storage products (especiallystarch) and excess nutrients, including phosphorus,in seeds, stems and other organs. Transfer of TP fromwater to the accumulating biomass of macrophyteshas been demonstrated by van Donk et al. (1993)and Meijer et al. (1994). This trait is invoked in thedeliberate use of macrophytes in small-scale waste-


water recovery (see Jewell, 1994, for an inspiringreview). The sites of storage within the plant may becrucial, for the annual die-back of vegetative tissueand its progressive disintegration, through hetero-trophic debris-detritus pathways, seems likely to leaknutrients back to the water. In cases of mildeutrophication, this can prove to be a significanteffect, at least in the short term, as macrophytic rootnetworks draw nutrients from the substratum andultimately release them into the water (Carpenter,1980, 1981). In mitigation, most of this detritalmaterial is processed in the sediments (van Donk et

al., 1993) and subject to retention there (see below).Thus, the generality about the role of macrophytesin the phosphorus dynamics of water bodies is that avariety of effects can be detected, ranging from netnutrient retention to a small, but often significant,net release (see review of Scheffer, 1998). In point offact, there is a much more consistent and predictablerelationship among case studies between the fluctu-ations in total nitrogen levels and relative macro-phytic cover. The behaviour of phosphorus inrelation to the presence or absence of macrophytes isclearly influenced by other complicating factors.

(4) Phosphorus diagenesis in bottomsediments

In accreting sediments where there are no higherplants, the sedimentary flux of biogenic materials issubject to diagenetic processing. The processes arecomplex and not altogether understood. The simplercases concern the oxic breakdown of the biogenicmaterials. Jewell & McCarty (1971) have given anauthoritative review of the organic breakdown ofsedimenting algae. A proportion (roughly one-third)of the cell carbon is typically very labile and subjectto biochemical oxidation within a day or so. Asecond fraction (also approximately one-third of theoriginal) decomposes rather more slowly (requiringweeks to months), while the remainder, mainlycellulose and skeletal derivatives, may be consideredrecalcitrant, if not altogether inert. Primary-biomasscarbon becomes the carbon of the decomposer andthe carbon of the first decomposer becomes thecarbon of the second decomposer, and so on.Maintenance and transfer costs see more and moreorganic carbon lost as carbon dioxide: even de-composer food chains comply with the laws ofenergy and lose biomass at each step! The relevantconsequence of progressive carbon oxidation is that,through stoichiometric conservatism, each microbialre-constitution acquires a surfeit of phosphorus (and

other elements) which is, thus, steadily dispersed asa consequence of the diagenetic catabolism. Thesame applies to the microbes heterotrophicallyoxidising the other sources of biogenic debris. Themaintenance of decomposer biomass occurs at theexpense of organic carbon in the system but to thedevelopment of a potentially recyclable excess ofphosphorus.

(5) Phosphorus-binding capacity ofsediments

This phosphorus may be liberated as organic P or asorthophosphate, which diffuses from around theimmediate sites of its regeneration. In fresh, activelymetabolising sediment, this inevitably means theinterstitial pore water, where it may accumulateas ‘ interstitial molybdate-reactive phosphorus’(IMRP; see Table 1). Further movement of thisdissolved phosphorus is governed by the laws ofdiffusion.

However, it is not inevitable that phosphateregenerated from sedimented biomass will remain insolution for it is open to chemical interaction withother sediment components. Ligand exchange at thesurfaces of metal oxides and hydroxides and of clayminerals, can result in the rapid sorption of newlyliberated phosphate groups from the ‘readily bio-available ’ (SRP) to the ‘conditionally available ’(NH

%Cl-P, NaOH-rP) fraction. Too few empirical

or experimental data exist to substantiate a generalpredictive model. In most of the fresh lake sedimentsstudied by Davies (1997), the largest single fractionof the phosphorus was ferric-bound; this fractionaccounted for over 50% of the total mass ofphosphorus present. Ferric hydroxide appears tohave an overriding affinity for phosphate ions insediments, just as it does in wet soils. Some degree ofproportionality between the phosphorus returned towater from oxic sediments and their TP:Fe ratio hasbeen observed by Jensen et al. (1992) and by van derMolen & Boers (1994). Significantly, for a subset oflakes where sediment TP:Fe is less than 0.1, thecorrelation is weak. This suggests that the iron inthese types of sediment is capable of immobilising upto one tenth its own mass of phosphorus. Thedynamics of sorption and desorption suggest that theequilibrium concentration increases with P-satu-ration. There seems to be no chemical reason whyTP:Fe should not reach 0.5 before significantquantities of excess phosphate are able to remain insolution.

The stability of calcium carbonate and apatites


engenders a view that calcareous lakes are effectivein immobilising free phosphate but the opposite isfrequently the case : retention of sediment-regen-erated phosphate by calcareous lake sediments hasnot been demonstrated, except where phosphate andcalcium salt are co-precipitated (House, 1990;Grobbelaar & House, 1995). Certainly, of the nativesediments investigated by Davies (1997), the lowestrelative phosphorus contents came from the twomost calcareous lakes. In one of these (the un-productive Malham Tarn; bicarbonate alkalinity2.3 mequiv l−"), where the elemental phosphorusaccounted for less than 1 mg P (g dry mass)−", thepoverty was possibly attributable to a chronicphosphorus deficiency in the entire system. In theother (the highly eutrophic White Mere; bicar-bonate alkalinity, 1.52 mequiv l−"), the sedimentphosphorus content was still relatively low, at only1.6 mg P (g dry mass)−", despite the SRP con-centration of the lake water being in the range1.8–2.4 g P m−$.

Davies (1997) went further in devising an assaytechnique for estimating the binding capacity ofnative sediments. Her experiments involved theexperimental exposure of lake sediments to preparedstandard solutions of orthophosphate (giving finalconcentrations of between 3.3 and 16.7 g P m−$) anddetermining how much was bound into whichfractions. In every case, she showed phosphatebinding began almost immediately, with saturationcomplete generally within 10 h or so. There were,nevertheless, major between-lake differences in thequantities sorbed. Deep-water sediments from fiveLake District lakes, containing from 1.5 to 5.1 mg P(g dry mass)−", all absorbed further phosphorus,principally into the NaOH-rP fraction, between 0.4and 3.2 mg P (g dry mass)−", consistent with ironbinding. In contrast, the HCL-rP fraction increasedby less than 0.05 mg P (g dry mass)−". The HCl-rPfraction in Malham Tarn increased by a similarmargin, though this was greater than the increase inthe NaOH-rP. The loose-bound phosphorus(NH

%Cl-P) increased in all experiments, broadly in

proportion to the concentration of added phos-phorus, but most in the Malham Tarn sedimentwhere up to 0.6 mg P (g dry mass)−" was sorbed. Nomeasurable sorption into the NaOH-rP or HCl-rPfractions of the White Mere sediment occurred,justifying the supposition that its phosphate-bindingcapacity is already wholly saturated. There is almostno sediment immobilisation of phosphorus. DP isthus able to exchange, within the limits of de-composition and diffusive transport, between the

superficial lake sediments and the water columnabove them.

(6) Redox- and pH-sensitive transformations

Elsewhere, the release of phosphorus back to thewater is dependent on the extent to which thesediment binding capacity might be altered, usuallyas a consequence of changes in pH or redox. It is self-evident that the sensitivity increases as a function ofphosphorus-richness of the system, thus of its primaryproductivity (to which alkalinity responds) and ofthe autochthonous organic carbon that may berecruited to the sediment (and to the oxidativedemands of which, the redox potential responds).

The most familiar (and, usually, the most signifi-cant) source of released phosphorus is the iron-bound fraction (strictly, the citrate-dithionite-extractable phosphorus or Na

#S#O

%-P but which

generally dominates the NaOH-rP fraction identi-fied in many analytical protocols). According to theEinsele-Mortimer model (see Section II.1), the onsetof low redox (!200 mV) in or close to organicbottom sediments leads to the reduction of amorph-ous ferric hydroxide to soluble ferrous and theliberation of sorbed and occluded orthophosphateions into solution.

Some of these will escape across the surface-waterinterface. If the redox of the water is also low, as isthe case in the summer hypolimnia of productivelakes, orthophosphate ions are released directlyinto the water. The SRP concentrations measuredin anoxic hypolimnia can be impressive (0.1–1 g P m−$) ; understandably, limnologists are con-cerned lest the renewed biosupportive potential inthe water column is realised. Some of their fear ismisplaced because, wherever the redox allows, ironre-oxidises, precipitates as ferric hydroxide and onceagain scavenges large quantities of phosphate fromsolution. It is not the potential bioavailability (in thesense of usability) of redox-sensitive phosphorus toaquatic primary producers that is in doubt so muchas its accessibility : for most phytoplankton, redox-mobilised phosphate is a remote resource, unless itcan be taken up mainly in the anoxic environmentswhere it is generated. For instance, it could be asource of phosphorus to anoxia-tolerant benthicresting cells and propagules. However, for otherplanktonic algae, the periodic solution and reconsti-tution of ferric hydroxide could make phosphorusless, rather than more, available to vegetative plantproduction in the upper, insolated water column. Inthe experiments of Davies (1997) and in cited


investigations of Cranwell et al. (1994) of thephosphorus-binding capacities of the surface sedi-ments of a selection of English lakes, a powerfulcorrelation was found between the quantity ofphosphorus that could be further assimilated into theNaOH-rP fraction and the amount of phosphorusbound in the fraction in the original sample.Apparently, the sediments already holding the mostbound phosphorus were the ones most able to bindmore. Frequent redox transformation of the sedimentiron evidently refreshes its phosphate-sorption pro-perties.

Raising alkalinity (i.e., increasing the concen-tration of hydroxyls) pushes the exchange equi-librium in the direction of displacement of immobil-ised phosphate by hydroxyls from the hydroxide(Drake & Heaney, 1987). The dynamics of thissorption-desorption pseudo-equilibrium are not wellcharacterised, so it is difficult to be categorical aboutjust how much phosphate is immobilised as solid-phase PP and how much is in solution. Although thesolubility products for ferric hydroxide and ferricphosphate in oxic, chemically neutral water are notprecisely known, it is to be supposed that they arewell towards the solid phase. Extreme dilution ofaqueous phosphate might be expected to favour de-sorption of phosphate. However, experimental evi-dence suggests that the proportion of sorbed phos-phorus that is released into deionised water is leastfrom sediments to which the greatest relativeamounts of phosphate are sorbed (Williams, Syers &Harris, 1970; Bostro$ m & Pettersson, 1982; Davies,1997). Some direct measurements of the SRPconcentrations in the interstitial waters of aerobic,non P-saturated, lake sediments suggest a range30–45 mg P m−$ (Reynolds, 1997) but the generalityof the statement is untested. Interference by (e.g.)silicate ions also influences desorption (Tuominen et

al., 1998).

(7) Whole-lake phosphorus budgets

Suppositions about the free passage of phosphorusbetween sediments and deep water are not alwayseasy to substantiate, neither is it always the case thatphosphorus released into deep water from low-redoxsediments plays a significant role in the carboneconomy of the lakes in which it occurs (Reynolds,1992). If the redox falls sufficiently low (approxi-mately ®100 mV) for sulphate ions to be reduced tosulphide, ferrous sulphide will be precipitated and,with it, the potential for the oxic binding of theliberated phosphate is removed (discussion in

Scheffer, 1998, p. 56). In lakes which are alreadyrich in phosphorus, such as the iron-poor, evaporite-rich kataglacial lakes of Shropshire, U.K. (Reynolds,1979), concentrations of orthophosphate phosphorus(1–2 g P m−$), far in excess of the demands ofplanktonic autotrophs, characterise the water col-umn throughout the year.

It becomes clear that there is no simple rule aboutwhether a lake sediment is a net source of bio-available phosphorus or whether it is a net sink, or,supposing that interactive transport of biologicallyavailable fractions (BAP) occurs, about the directionof the net fluxes at any given point in time. Thoughundoubtedly complex, it should be possible toformulate some testable hypotheses or a generalisedword model about the recycling and reuse ofphosphorus within whole lakes.

The basic reference point remains the annualexternal load of usable phosphorus, Σ

y(SRP). The

simplest compensation is that the external load isprecisely balanced by the total mass of phosphorusthat is removed annually through the outflow(TP

OUT) plus the annual increment to the sediment

arising through biologically-mediated trans-formations (δP

sed). Thus,

Σy(SRP)¯TP

OUTδP

sed(6)

As the organic component of the sediment is oxidised,phosphate is sorbed at available binding sites on theclay minerals, oxides and, especially, ferric hy-droxide, becoming immobilised in the ‘conditionallyavailable ’ fraction. Pending complete burial, modestbiogenic fluxes in P-deficient, iron-rich lakes are thuslikely to be retained in lake sediments.

The mechanisms by which conditionally availableP becomes free BAP or is diffused into the watercolumn have been characterised. The critical quan-tity is the proportion which the sediment retains,which, to a large measure, must be related to itsresidual binding capacity. As the data from Shrop-shire Meres indicate, binding capacity can besaturated. The excess stays in solution, initially asIMRP, or is leaked, by diffusion or with theassistance of physical or biological ‘ turbation’. Thereis no longer a chemical barrier preventing re-use ofrecycled phosphorus in the water column.

Supposing the generality of this simple model, theability of a given sediment ‘ to release ’ phosphorusshould be broadly predictable from the measurementof its instantaneous residual binding capacity. Theextent to which the capacity is filled by the biogenicinflux and the liability of the sediment to mechanicaldisruption are similarly indicative of system sen-


sitivity to recycling. The critical values are subject toa great deal of between-site variability but eachsystem might be supposed to be subject to some site-specific threshold value, [P]

k, below which sediments

might be expected to retain phosphorus. On thebasis of a series of case analyses, Sas (1989) suggestedthat, as a general guide to its magnitude, [P]

khas a

value of around 1 mg P (g sediment dry mass)−".There is undoubtedly scope for considerable be-tween-site variation in this threshold. Davies (1997)found sorptive capacities of 5–7 mg P (g sedimentdry mass)−" in accreting deep-water, iron-richsediments. Conversely, [P]

kseems likely to be least in

the iron-deficient or calcareous sediments of slow-flushing, phosphate-rich lakes. The greatest op-portunity for the release and re-use of sedimentaryphosphorus might depend most upon a combinationof high phosphorus concentration, weak bindingcapacity and frequent exposure to physical dis-turbance as a consequence of shallow position oroverstocking by benthivorous fish (especially cypri-nids : Lamarra, 1975).

(8) Some case studies

A selection of site-specific phosphorus budgetssupports the above hypothesised generalisations.Starting with Loughgarve, a shallow, upland lake inNorthern Ireland receiving a modest, mostly atmos-pheric P input (approximately 0.1 g P m−#) andexperiencing rapid hydraulic flushing, Gibson(1997) has shown that very little phosphorus isaccumulated in the sediment at all. In the larger,deeper and hydraulically more retentive LoughErne, the catchment contributes most of the annualphosphorus load (approximately 1.9 g m−#), some35% of which is retained in the sediments, whencereturns to the water are negligible. In Lough Neagh,a relatively large, shallow and retentive lake (re-tention time 1.25 years), most of the annual P load(1.2 g P m−#) sediments with the spring bloomdiatom bloom, although a proportion of this issubsequently released back into the water. There itsupports further algal growth and another depo-sition, representing some 30% of the annual load(Gibson, 1997).

Windermere, UK, is a deeper (maximum depth,64 m), glacial ribbon lake which receives a catch-ment-derived background load (0.4 g P m−# year−")and, until recently, an input of secondarily-treatedsewage effluent (generating a maximum further loadof C 0.9 g P m−# year−"). Of this, 85% settled

permanently to the lake bottom, without anydetectable subsequent release back into the watercolumn (Reynolds, 1999). In the nearby Bassen-thwaite Lake, however, which resembles Winder-mere in the chemistry and volume of its annualhydraulic load, as well as in receiving a secondarily-treated sewage effluent (maximum TP load,1.6 g P m−# year−") but differs substantially in itsbathymetry (mean depth: 5.5 m against 21.3 m)and in its sediment phosphorus content [1.5–2.0against approximately 4 mg P (g dry sediment)−" :Davies 1997], the annual net biomass productionconsumed up to four times the annual P load; activephosphorus recycling from shallow sediments ispowerfully implied. In the phosphorus budgetssolved for two, uniformly deep (approximately11.5 m) large limnetic enclosures (no outflows!),installed in another small lake of the Windermerecatchment (Reynolds, 1996), almost no recycling ofphosphorus loads of between 0.3 and 1.9 g P m−#

year−" could be detected. A graded bottom, from 2.5to 13 m depth, in an otherwise similar enclosure inthe same lake permitted a production yield of up totwice the capacity of the annual phosphorus load.The role of wind-induced resuspension in refreshingP availability in the trophogenic was demonstrated.These cases emphasise the role of retention time (inallowing assimilation of the load in the first place)and water depth in providing the opportunities forre-use of sedimented phosphorus.

In Lake Michigan, one of the world’s oligotrophicgreat lakes, concern has centred on the effect ofenriching riverine inflows on the water quality in theshallower areas of the lake. Green Bay has been onesuch area, where 70% of the annual phosphorusload is received through one river alone. Klump et al.(1997) determined that most of this sedimentedwithin the confines of the Bay (approximately0.02 g P m−# year−") where, subject only to ‘ focus-ing’, 70–80% of this was retained and buried withinjust 20% of its area. Here, the sediment phosphoruscontent was found to vary in the range, 0.15–2.2 mg(g dry sediment)−". For Lake Mendota, Wisconsin(39.9 km#, mean depth 12.7 m), Soranno, Carpenter& Lathrop (1997) measured annual aggregates ofphosphorus sedimentation to be between 2.9 and4.6 P m−# but, to balance these against knownexternal loads, they had to assume that stormresuspension and entrainment recycled sedimentaryphosphorus at up to ten times the external loadingrate. In the Eau Galle Reservoir, Wisconsin, havingan extensive shallow area (! 4 m deep), James &Barko (1997) found that the external phosphorus


load was retained but for an aggregate annualbiomass assembly some 24% greater than the annualload could sustain theoretically. Hamilton &Mitchell (1997) were able to show that in shallowWestern Australian lakes, the phosphorus ‘releases ’were driven mostly by wave-induced shear. Mostimpressive of all is the internal load to Søbygaard,Denmark (Søndergaard, Kristensen & Jeppesen,1993) which, despite the imposition of a sharpreduction in external loading (from approximately30 to ! 5 g P m−# year−"), the lake sediment con-tinued to supply phosphorus to the water at rates of3–8 g P m−# year−").

The Søbygaard case is one of the better docu-mented instances of a shallow enriched lake (forseveral others, see the review by Sas, 1989) where,once the external load had been lowered, δP

sed

became a negative term (i.e. indicative of net internalP loading), at least until such time as the sedimentphosphorus concentration falls back below [P]

k. It is

deduced that, besides supplying the ‘resilience’ oflakes to lowered external phosphorus loading, thisfree exchange between P-saturated sediment occursanyway, at least while the external loading is stillheavy and unregulated, and especially so when it isassisted by mechanical disturbance (Peters &Cattaneo, 1984; Kallio, 1994). In eutrophied, soft-bottomed river channels, almost continuous ex-changes are facilitated. Where what may alreadyamount to an abundant, freely available resource toautotrophs is present in the water, it is taken up byphytoplankton, transported as BP to the sedimentand whence, at concentrations much exceeding [P]

k,

it is freely returned to the river, albeit a little furtherdownstream, by a sort of saltation process(Grobbelaar & House, 1995). As a corollary, it mustbe supposed that in shallow, dynamic waters whereorthophosphate levels in the water regularly exceedthe [P]

kthreshold of the sediment, the excess

phosphorus moves freely and near-indefinitelyamong the system’s other components.

It should be emphasised, however, that it is nomore than the superficial veneer of sediments that isinvolved in phosphorus exchanges. The fears thatthe entire sediment P might somehow be releasedfrom the bottom mud are misplaced. Essentially,there are only two routes by which liberatedsediment phosphorus re-enters the water column.One is through molecular diffusion of IMRP. Thenumber of moles of solute (n) passing across a unit ofsediment surface (b) in time, t, is a function of thesolute concentration (C), the concentration gradi-ent in the vertical direction (dC}dz) and D, the

coefficient of molecular diffusion (approximately10−* m# s−") :

n¯ b D (dC}dz) t. (7)

Thus, a difference of up to 150 mg P m−$ (i.e.,5¬10−$ mol m−$) across the mud-water interfacewould sustain a flux of about 5¬10−& mol m−#

day−". Release rates of this order (1.5 mg P m−#

day−") have been measured from very P-richsediments (Ahlgren, 1977; Sonzogni et al., 1977;Keizer & Sinke, 1992). The formulation also predictsthat if the concentration difference is to remainconstant, its position must move downwards into thesediment: in theory, this could be up to 10−".& m, or30 mm, in a year. This distance then defines thethickness of the ‘active layer ’ of sediment: a uniformsediment would have exhausted its unbound IMRPthrough that depth within one year.

The second route is through mechanical dis-turbance. At its mildest, convective water circulationis sufficient to maintain the favourable diffusiongradient ( James & Barko, 1991). Mechanical mixingof the sediment surface, sufficient to resuspendsedimented material and entrain interstitial water,requires the application of shear energy, generatedby wind or waves and transmitted through the watercolumn. A strong wind, say of 20 m s−" andintroducing a mechanical energy flux of 1.6¬10−# W m−# at the surface, is propagated downwardsto extinction in approximately 180 m of a deep lake,at a dissipation rate of C 2.2¬10−( m# s−$. The sameenergy flux in a shallow (C 9 m) lake would have tobe dissipated 20 times faster (4.4¬10−' m# s−$) andwould penetrate the fine sediments. Nixon (1988)compared the disturbance wrought by differentstress levels on different textures and found that thedepth of fine sediment required to dissipate thepenetrating stress was generally less than 40 mm andalways less than 50 mm. Thus, rather less than thetop 50 mm of mud surface can ever be considered‘active sediment’.

(9) Internal phosphorus recycling: theemerging overview

From the time that Einsele and Mortimer persuadedlimnologists that sediments were integral to la-custrine processes, the science has struggled withconfusing messages about the complex set of factorsgoverning the movement of phosphorus between thewater column and its lower surface. The most recentwork confirms that bioavailable phosphorus isreleased from oxic as well as anoxic sediments and


that it does not emanate only from deep sediments(Ahlgren, 1977; Riley & Prepas, 1984), nor only ineutrophic systems (Twinch & Peters, 1984). Theemerging paradigm is that whether, when and howmuch biologically available phosphorus is returnedto the water is, firstly, a function of the sedimentaryflux (in the long term, there cannot be more returnedfrom the 30–40 mm active layer than has beenrecruited to it) and, then, a function of theoxide}hydroxyl phosphate-binding sites availablewithin the active layer. The scarcer these areoriginally (the less clay mineral, the less aluminiumand the less iron present) or the more they arealready occupied by orthophosphate radicals, thenthe lesser will be the inactivation of the ortho-phosphate ions and the greater the potential for their‘ release ’ into the water column. Moreover, the morethat the sediments are exposed to wave or shearstress and bioturbation, then the more likely it is thatphosphorus will be recycled freely and be re-usedbiologically.

In the context of establishing the potentialbioavailability of the various fractions of the totalphosphorus reaching fresh waters, it may be re-assuring that the ‘readily-available ’ fractions gen-erally represent a small part of the total phosphorusload emanating from catchments. This must betempered by the magnifying impacts of internalrecycling. Compared to the approximate biomassyield (see Section III.2 : 1 mg plankton chlorophyllper mg bioavailable P), internal phosphorus re-cycling and re-use means that the temporal pro-ductive-yield capacity of phosphorus is almostinfinitely variable. It may be close to 1 (for deeplakes with retentive, binding sediments), rising to1–10 mg aggregated chlorophyll product per mg Pload (for P-enriched, shallow calcareous lakes subjectto frequent sediment entrainment) and, at least intheory, many more times greater than that (Kallio,1994). The appreciation that really quite a smallamount of phosphorus can support an ongoingbiomass and, over time, a substantial aggregate, hasto be adopted into present general understanding.

V. CATCHMENT SOURCES OF PHOSPHORUS

(1) Phosphorus in soils

In this section, we work back to the sources ofphosphorus in hydraulic catchments and to theprocesses yielding significant fluxes of BAP to water.Interestingly, in spite of its relative abundance in theearth’s crust, indications of the TP content of most

topsoils (0.05–1.1 g kg−" : Brady, 1990; 10–300 g P m−# : Stee!n, 1997) suggest they are naturallyrather P deficient. Moreover, infertility is com-pounded by the paucity of labile P; most resides inthe scarcely bioavailable fractions, which are gen-erally incapable of meeting the maximal demands ofrapid plant uptake. On agriculturally unimprovedland, as in water, the fluxes of bioavailable forms ofphosphorus are critical to biological productivity.Concentrations of dissolved phosphate in soil waterare generally modest (usually ! 0±2 g P m−$ : Stee!n,1997), so that natural vegetation and, especially,climactic forest take many centuries to develop andmature. Such phosphorus as these terrestrial ecosys-tems are able to amass tends to be retainedintrabiotically, either in living biomass or in mi-crobial reprocessing of dead remains. When scrub orforest is cleared for agriculture, fertility may beremoved with consequentially poor crop production.Agricultural yields cannot be expected to exceed theflux capacity of the relatively least available plantnutrient. In younger soils, at least, where other,more labile nutrients (like nitrogen) persist, phos-phorus certainly can be a yield-limiting nutrient.

The most common primary inorganic forms ofphosphorus in the soil are apatites (calcium phos-phate) and phosphates of iron (e.g. strengite) andaluminium (variscite). These are generally notdirectly bioavailable, pending slow weathering anddissolution as soil-water SRP. Plants will readilytake this up and assimilate it into biomass but theydo so in competition with the mineral bindingaffinity. In this way, most of the inorganic phos-phorus in developed soils is incorporated in ‘ secon-dary minerals ’ – hydrous sesquioxides, amorphousiron or aluminium oxides and hydroxides. None ofthese forms are necessarily any more exploitable byplants, save at the pace of sorption-desorptiondynamics.

That which is eventually incorporated into bio-mass can find its way back to the soil, perhaps in leaf-fall, plant death and decomposition or in the bodiesand wastes of consuming animals. The progressiveaccumulation of organic phosphorus compounds insoil – including inositols, phospholipids, fulvic andhumic acids – is often crucial to their continuedfertility. Yet even microbial mineralisation of thenatural organic pool is modest and protracted(generally, ! 1–2 g P m−# annually ; Brookes,Powlson & Jenkinson, 1984).

The fractionation of soil phosphorus is, thus,naturally quite stable or, at least, changes too slowlyto sustain economically viable agricultural pro-


duction. The improvement of soils for the raising ofcrops became an objective long before the finerpoints of their chemistries were understood. How-ever, the traditional application of organic compostsand manures is just one of the ways in which thepotential bioavailability of phosphorus in soils maybe augmented (Tunney et al., 1997). Modernapproaches to soil husbandry developed during theeighteenth-century with the introduction of seed-drilling, crop rotation and, significantly, the use ofregular dressings with manure and bone meal (seeJohnston & Poulton, 1992, for an interpretativereview). Eventually, the importance of providing agood supply of water-soluble phosphates was seen tobe the key to improving agricultural yields. More-over, because soils tend to retain phosphorus readily,application in excess was thought to be desirable.

The phosphorus chemistry of soils is now known tobe exceedingly complex, involving numerous in-organic compounds and organic derivatives (seeSharpley & Rekolainen, 1997). For most of thepresent century, the use of inorganic compoundedfertilisers has supported remarkable rises in the area-specific productivity of soils reclaimed for agriculturefrom former grasslands and woodlands, many ofwhich would otherwise show symptoms of yield-limiting phosphorus levels and would be capable ofsustaining only poor crops or poor-quality grazing.Tree-dominated systems are more bioretentive thanthe simplified organisations imposed in their place,while the replacement of fungi by bacteria as themain agent of heterotrophic recycling may also haveinfluenced the phosphate-application rates deemedto be required for soil improvement.

(2) Fertiliser application strategies

In purely agricultural terms, the ostensible purposeof applying phosphatic fertiliser is to deliver sufficientbioavailable nutrient to fulfil the potential growthrequirement of the crop. To bias the outcome ofcompetition with chemical binding sites, it isimportant to deliver the dose adequately close toplant roots (within approximately 2 mm: Nye &Tinker, 1977). Ideally, application of phosphorusfertiliser supplements the ability of the existing BAPto deliver sufficient orthophosphate for plants todraw in enough to sustain an acceptable crop yieldor conversion to animal product (Morgan, 1997). Inpractice, the outcome is attained by adding phos-phate in excess.

Good agronomic practice, just as much as concernover possible phosphorus leakage from soils, requires

sound understanding and reliable analytical de-termination of the dynamics of phosphorus inagriculture. Good data sets on the phosphoruscontent of soils and of the crops harvested therefromare widely available. As stated earlier (see sectionII.2), the techniques for fractionating and assayingthe phosphorus were devised originally by soilscientists (Deans, 1938; Chang & Jackson, 1957).Although invoking analogous analytical principlesto hydrochemistry, the sampling methods and thesensitivity targets of soil chemists are different andthe nomenclature is different. ‘Olsen P’ (alkali-extractable fraction removed in 0.5 m sodium bi-carbonate solution at pH 8.5: Olsen et al., 1954) is ameasure of the iron- and aluminium-bound phos-phorus and is taken in many countries to be the bestmeasure of what is potentially available to crops.Elsewhere, ‘Morgan P’ (that extractable in 10%sodium acetate : Morgan, 1941) is preferred as ameasure of the more readily exchangeable P. Strongacids, acting on calcium-bound fractions, are used inthe determination of ‘Bray P’ and ‘Mehlich P’.Water-soluble, CaCl

#-soluble and total-P fractions

are distinguished in some assays (Sibbesen &Sharpley, 1997).

Several recent reviews help to establish theintensity of P application necessary to raise soilfertility to a point where areal yields of arable cropsor of herbivore meat or milk meet conventionalstandards of agricultural efficiency. The harvest ofcereals, such as wheat and barley (in which virtuallythe whole annual plant production is removed fromthe ground, may deplete the soil reserve by 3 g P m−#

year−" (Tunney et al.,1997). The average net removalof phosphorus from grazing land, as milk or meat, isnearer to 1 g P m−# year−". Yields of root vegetables(beets, potatoes) are saturable in soils containing inexcess of 25 mg Olsen P kg−" (Johnston & Poulton,1997a, b). Supposing a plough depth of 0.2 m and anaverage density for ploughed soil of approximately1000 kg m−$, the mass of relevant topsoil may beapproximated as 200 kg m−#. Then 5 g availableP m−# year−" may be deduced to be a more thanadequate application for the maintenance of highfertility. In fact, the applications recommended invarious countries range between 20 and 100 kg ha−"

(2–10 g P m−# year−") for low-P soils and betweenzero and 40 kg ha−" (i.e. up to 4 g P m−# year−") insoils already containing " 40 mg Olsen P kg−" (datasummarised in Tunney et al., 1997).

The guidelines are directed towards the deliberateaccumulation of bioavailable P in the soil. Duringthe development of an agricultural soil, the practice


of adding excess phosphorus provides a small risk ofgenerating an excess of soluble and, hence, leachableP. However, continuation of the process raises theprobability of export of P other than in targetbiomass. It may not be necessary to maintain eventhis level of phosphorus availability. Tunney et al.(1997) refer to the results of plot trials in which threecuts of grass for silage each year, yielding an average12 Mg dry matter ha−", cropped roundly40 kg P ha−" (4 g P m−# year−"). This was repeatedin each of the next eight years (during which a totalof 35 g P m−# was harvested) without significantreduction in annual crop and without furtheraddition of phosphate. The phosphorus accumulatedin the soil during the preceding forty years proved tobe adequately bioavailable.

(3) Other catchment phosphorus sources

Any supposition that the retention of inorganicphosphates in agriculturally improved soils is com-plete must be rejected. The perspective that isneeded is the magnitude of the diffuse outputs fromagricultural land in comparison with other diffusesources and the point sources arising from increas-ingly urbanised and industrialised human popu-lations.

The yields from urban ‘agglomerations ’ (theterminology of the EU Urban Waste Water Di-rective) are well characterised (broadly as quantifiedin Section I.4 above). Other potential sources ofexportable phosphorus arise from rough or ‘un-improved’ grazing land. With labile phosphoruslevels towards the lower end of the range cited byStee!n (1997; see Section V.1 above), the SRP exportin drainage is also trivial. Conversely, such phos-phorus as is exported will be largely in particulateform and only scarcely bioavailable. The smalldatabase of estimates of phosphorus exports from(largely) unimproved catchments to lakes, includedin Table 2, concur with this deduction. Even onimproved soils, where the area-specific TP yieldsmay be an order of magnitude larger, by far thelargest fraction is still represented by eroded soilparticles. Huttula (1994) showed that most of thephosphorus transported to the Finnish lake,Pyha$ ja$ rvi, was bound to suspended soil particles,eroded from its catchment. Supposing it to haveoriginated mainly from fertiliser dressings, thebinding of this phosphorus to iron or aluminiumoxides, or its combination with calcium, effectivelyimpairs its bioavailability.

Organically bound phosphorus is generally also

particulate. Plant remains, leaf litter and fragmentsyield potentially available breakdown productsthough, as acknowledged (Section II), these may bequickly immobilised too. The same should apply tothe phosphate content of farmyard manure andslurry dressings, although the practice of applyingthese liberally over the surface of grassland forpasture or fodder harvest may well separate the sitesof phosphorus diagenesis from the sites of mineralsorption sufficiently to bias its assimilation by thegrass crop. Manuring fields is an efficient and a cost-effective means of restoring nutrients to the land butthe practice is vulnerable to the loss of bioavailablephosphorus in run-off, especially from slopes duringwet or frosty weather. Local controls on applicationrates and times are imposed in some countries (e.g.Switzerland) with the express purpose of minimisingeutrophication.

(4) Transfer pathways from soils to surfacewaters

Several recent reviews (Sharpley et al., 1995;Haygarth, 1997; Morgan, 1997; Sharpley & Reko-lainen, 1997) have addressed the comparativeimpacts on fresh waters of what are, in all prob-ability, relatively large transfers of chemically immo-bilised phosphorus and of agronomically trivialleakages of dissolved and readily available SRPfractions. There is agreement that particulate phos-phorus (PP), including that sorbed to soil particlesand organic matter (dead leaves and other biogenicfragments), generally represents much the largestfraction transported from cultivated soils (as muchas 60–90% of the TP removed is in this form:Pietila$ inen & Rekolainen, 1991; Sharpley et al.,1992). Some of this can be carried as wind-blowndust, although the main avenue for particulates isunderstood to be in water flow. The particulateseroded and transported from well-managed per-manent grass (i.e. neither overgrazed nor ‘poached’)and maturing woodland are, however, minimal.Factors such as soil type, thickness, slope, the natureof past land-use and the present vegetation coverthus directly influence the scale of particulatephosphorus loads shed to water from specificcatchments and at specific times.

Because much of this PP is thought unlikely toinfluence the productivity of the receiving waters tonearly the same extent as do the SRP loadings, thelatter require closer consideration. Dissolved, col-loidal or fine-particulate phosphorus is transportedfrom the topsoil along three main hydrological


Table 2. Losses of total (TP) and soluble phosphorus (SRP) from various types of catchment, in mg P m−# year−"

TP SRP Reference

Mature temperate forest C 2 Hobbie & Likens (1973)Mature boreal forest 3–9 Ahl (1975)Boreal forest catchments 5–11 1–4 Pietila$ inen (1997)Cleared forest, igneous C 5 Dillon & Kirchner (1974)Cleared forest, sedimentary C 11 Dillon & Kirchner (1974)Cleared forest, volcanic 72 Dillon & Kirchner (1974)

Woodland, grassland (E. France) 1–30 Pommel & Dorioz (1997)Mixed upland (English Lake District) 24 5 Reynolds (1999)Mixed lowland (S. England) 340 22 Withers (1997)

Pasture (USA) 8–20 Johnston et al. (1965)Grass, unfertilised (New Zealand) 22–117 2–50 Sharpley & Rekolainen (1997)Grass, unfertilised (US) 40 10 Sharpley & Rekolainen (1997)Grass, upland, improved 70 23 Withers (1997)Grass, improved (N. Ireland) 0–180 Lennox et al. (1997)Grass, improved (New Zealand) 33–554 4–280 Sharpley & Rekolainen (1997)Grass (UK) 100 Hayward et al. (1993)Grass, grazed 127–460 5–14 Haith & Shoemaker (1987)Grass, improved (US) 970–1240 43–140 Sharpley & Rekolainen (1997)Grass, with low runoff 12 van der Molen et al. (1997)Grass, with high runoff 202 van der Molen et al. (1997)

Intensive arable (UK) 7–25 Cooke & Williams (1973)Intensive arable (Netherlands) 25 Kohlenbrander (1972)Arable, sandy soil 26 van der Molen et al. (1997)Arable, with waterlogging (US) 40–53 Schuman et al. (1973)Arable, with heavy erosion (US) 190 Johnston et al. (1965)Cultivated (E. France) 50–200 Pommel & Dorioz (1997)Cultivated (Finland) 77–170 15–41 Pietila$ inen (1997)Intensive arable (UK) 163–260 24–72 Withers (1997)

Wheat (US) 294 34 Haith & Shoemaker (1987)Wheat (Canada) 410 120 Sharpley & Rekolainen (1997)Maize (US) 60–160 20–100 Sharpley & Rekolainen (1997)Maize (US) 180–200 70–80 Sharpley & Rekolainen (1997)Maize (US) 1390–1770 10–20 Sharpley & Rekolainen (1997)

Eroding slopes, Mediterranean 200–1400 Chisci & Spallacci (1984)Vineyard 690–1020 Pommel & Dorioz (1997)

Groundwater leachates 50–150 Reynolds (1979)Yield from unmodifiedrainfall runoff

5–80 Hypothesised (see Section IV)

pathways : in surface run-off (overland flow); in sub-surface run-off (leachate, throughflow which re-mains separate from local water tables) ; or in sub-surface groundwater flow. This nomenclature wasproposed by Ryden, Syers & Harris (1973) and itsgeneral adoption has been advocated by Foy &Withers (1995).

Run-off refers to the exportable atmospheric pre-cipitation, net of evapotranspiration from the groundsurface, and which, by gravitation and aggregation,becomes the surface flow in streams and rivers.

Surface run-off (or overland flow) is that part of thetotal run-off that flows over the surface of the groundto reach discrete stream channels (including ditches)directly without penetration of the soil. By thisdefinition, it is rain water and snow melt andcontains only solutes leached from the atmosphere orfrom contact with the ground surface. It is generatedwhen the rate of surface accretion of water exceedsthe rate of its removal by percolation into theground. Realistically, surface run-off is promotedduring and immediately after heavy precipitation.


Discharge of surface run-off increases with theintensity of precipitation and with the relativesaturation of the soil but decreases with its porosity.The steeper is the slope of the surface, the more is thebias of drainage to surface run-off. The phosphorusremoved in surface run-off includes the DP}SRPfractions present in the original precipitation (typi-cally in the order of 10–100 mg P m−$ ; see SectionIV.1 above), together with any freshly dissolved orde-sorbed at the soil surface (such as from fertilisergranules). It also includes non-soluble particulatefractions entrained in the flow. Fine soil particles,plant litter and animal remains, as well as significantquantities of organic material, arising from manureand slurry applications, can be transported in thisway. Such run-off is prevalent over made surfaces(farmyard hardstanding, highways), whence looseparticles are also entrained. This can represent acocktail of organic and inorganic phosphorus, fromwhich solution and desorption may increase the BAPcontent. These processes can result in significantlocal exports of phosphorus (Haygarth, 1997).

Sub-surface run-off (or interflow) is that part of thetotal run-off which infiltrates the ground, movingdown and, often, laterally, usually in response togravity, reaching stream drainage without con-tacting a local water table. This water is open tomodification through solution, concentration, selec-tive uptake by plants and micro-organisms and itcarries the metabolic residues of biotic intervention.As plant uptake of SRP competes with unsaturatedchemical binding capacity, phosphorus moves fromone scarcely soluble form to another or to croppablebiomass. While readily assimilable phosphorus re-mains deficient, dissolved forms are removed veryefficiently from solution in the sub-surface soil water.Only when the phosphate-binding capacity ap-proaches saturation is free SRP likely to be abundantand to accumulate in plant biomass. Consequently,no great variation in the net balance of phosphorusexchanges or in the residual SRP concentrations ofthe percolating or throughflowing water can behypothesised simply on the basis that excess phos-phate is applied to the soil surface. Measurements ofthe SRP content of soil waters exceeding10 mg P m−$ are indicative of sorption and uptake ofthe phosphate present in the incoming rainwater,whereas 200 mg P m−$ is indicative of a fertile soil(Ryden et al., 1973; Sharpley & Menzel, 1987;Stee!n, 1997). It is possible for saturating fertiliserapplications to persist in solution in soil water and,thus, for dissolved phosphate to be exported intosubsurface drainage.

SRP in the sub-surface flow can be transportedonwards and over distances determined by the nextsoil horizons encountered. If abundant P-bindingcapacity exists a little deeper down, then phosphatewill be immobilised there. Soils typical of improvedpastureland seem to conform to the pattern recog-nised by Haygarth (1997), where, for horizonsgreater than 150–300 mm, mobile phosphorus isdeficient. However, shallow or sandy soils, ones onsteep hillsides or ones into which tile drainage isinserted are correspondingly more likely to yieldSRP-rich water to streams and rivers. The analyticaldata of van der Molen et al. (1997) indicate that sub-surface run-off to ditches draining phosphorus-saturated, sandy upper soils accounts for " 70%of the 0.1–4.0 kg P ha−" (0.1–0.4 g P m−#) exportedannually. The amount is agronomically trivial buthydrobiologically disturbing.

Groundwater run-off is that part of the total run-offwhich percolates through the main soil horizons toenter the ground water at the water table. Thiswater, too, gravitates slowly to constitute regionalgroundwater flow finding its way to stream channelsor lakes where the surface topography intercepts thewater table, perhaps many kilometres downstream.The importance of the groundwater pathway in theoverall mobility of phosphorus is not well established,although, in some instances, the concentrationthresholds of eutrophy (say, 35 mg P m−$) would bereadily exceeded. Work on the ground waterssupplying the kataglacial moraine lakes in Shrop-shire, UK (Reynolds, 1979) found SRP concentra-tions to be in the range 0.21–0.59 g P m−$. However,it is not clear that the phosphorus is derived fromsurface soil leachate so much as from the evaporite-rich subsoil.

(5) Phosphorus transfers at the catchmentscale

Unless it is possible to mount field campaigns ofsample collection and analysis, with an intensity intime as well as space, quantification of the diffuse-source transfers of phosphorus to water at thecatchment scale has to rely on simpler, statisticalapproaches. One of these invokes analysis of time-series of careful measurements of the various phos-phorus fractions at selected points in the recipientcatchment and their backcorrelation to knowncatchment activities. This can work reasonably wellif the land-use is fairly uniform and the point sourcesare adequately quantified, although multiple ac-tivities make the individual impacts difficult to


distinguish (Heathwaite, 1997; Lennox et al., 1997).The alternative is to concentrate upon particularland-usage categories across a range of individualcatchments so that the order of typical exports fromgeneric catchment categories can be approximatedand applied to the calculation of catchment-widebudgets (Harper, 1992). The trouble with this is thatno two catchments behave exactly alike, so thatapproximations have to carry a generous margin oferror. The better option is to combine the approachesand to seek general patterns from the correlation ofloading trends with altered land-use practices (seeFoy & Lennox, 2000, for an apposite case study).

In spite of the wide variability in catchmentprocesses and questionable accuracy and reliabilityof the available data, the tabulation of literature-derived phosphorus exports (Table 2) does revealsome general trends about their nature and scale.Owing to detail differences in the chemical deter-minands chosen in the studies, the distinction ismade only between total (TP) and soluble molyb-date-reactive phosphorus (SRP), where such dataare given. The tabulation confirms that the leastarea-specific phosphorus exports emanate from ma-ture forested catchments (2–11 mg m−# year−"),which remove and store atmospheric P inputs (5–80 mg m−# year−"). The sequestration and retentionin the plant biomass is further emphasised by theseverity of nutrient deficiency in soils cleared of long-standing forest for agriculture (Ewel, 1983; Finegan,1996). These citations refer to the tropics and thecrucial nutrient in question was nitrogen but theecological principle of resource conservation inadvanced successional stages is strongly upheld.Certainly, once forest cover is removed, the dynamicsof phosphorus movements are increased (Hobbie &Likens, 1973; Omernik, 1977).

We may hypothesise that the pre-settlement, pre-agricultural, temperate-forest landscapes exportedvery little SRP and that receiving rivers and lakeswere, accordingly, typically deficient in bioavailablephosphorus and, hence, always oligotrophic incharacter. The corollary, that removal of forest andthe superimposition of agriculture necessarily in-creases the export of phosphorus to water and, so,fuels eutrophication, is frequently advanced. Thereis sometimes good palaeolimnological evidence forthis (Reynolds, 1998, discusses the case of Winder-mere). A more reasonable deduction invokes thepartitioning of the phosphorus during the devel-opment of the soil. Labile orthophosphate (SRP)remains a scarce or transient resource in unimprovedsoils and many improved ones too (see Section V.4,

above). In part, calcite fulfils this binding role inchalk soils ; hydroxy-metal phosphates of aluminium,iron or manganese form predominantly in acidicsoils. While binding sites remain in the soil andsubsoil, losses of dissolved phosphorus to drainagewater are substantially minimised or confined toepisodic surface run-off events. Creation of pasturefrom woodland is not necessarily the trigger forlarge-scale export of phosphorus (Johnston et al.,1965; Pommel & Dorioz, 1997; Sharpley & Reko-lainen, 1997) but repeated fertiliser applicationsmay, in time, lead to much-accelerated TP losses(Table 2).

The entries in Table 2 also attest to acceleratedexports of total phosphorus (TP) from arablecroplands. Moreover, from the instances where datafor SRP losses are given, the proportion of the exportthat may be readily bioavailable in aquatic habitatsis itself a conspicuous variable. The challengeremains to distinguish among three possible explan-ations. Either (i) the primary mechanism is theremoval of phosphorus bound to eroding soilparticles (TP moderate to high but SRP low); or (ii)the transport owes to enhancement of the labilefraction in the soil (TP variable but SRP propor-tionally dominant) ; or (iii) there is a failure of thesoil to immobilise phosphorus (in which case, the TPand SRP exports are quantitatively similar).

The potential for soil erosion is increased bytopsoil ploughing in preparation for monospecificseeding. Frost weathering is often a deliberate step ingaining soil friability but the diminution in particlesize increases particle entrainability in surface run-off. Rain water is the primary medium of phosphorustransfer and its efficiency is enhanced when it flowsacross the surface, scouring fine materials as it doesso. Removal of particulates from arable land is, mostof all, a function of soil type, precipitation intensityand surface gradient. Rainfall exceeding 5 mm h−" issufficiently intense to cause significant erosion(Boardman & Robinson, 1985) but lesser intensitiesfalling on wet soils can promote surface flowsadequate to entrain soil particles into streamdrainage (Evans & Northcliff, 1978), as well asbiogenic litter and organic matter as described inSection V.4, above.

Other things being equal, the transport of phos-phorus-containing particulates is greater from cul-tivated ground than from grassland (Pietila$ inen,1997; see also Table 2) and from waterloggedground than from areas of freely draining sandy soils(Haygarth, 1997). In this case, percolation increasesthe opportunity for the chemical immobilisation or


biotic assimilation of the phosphorus (Schoumans &Breeuwsma, 1997). Interestingly, a good clay con-tent represents added sorptive capacity, but this iscountered when its drainage is impeded and sus-ceptibility to waterlogging is increased and by theonset of reducing conditions. Conversely, the steeperthe slope, the greater is the tendency for theformation of discrete drainage channels, with aconsequent enhancement of surface run-off of par-ticulate P. The point is emphasised again that verylittle of this is readily available to aquatic algae(Ekholm, 1994).

Instances of accelerated SRP export from catch-ments, not otherwise attributable to point sourcesare less common. While percolation of rain waterideally leads to deep phosphorus retention, theapplication of mineral fertiliser is intended to raise orprolong bioavailability in the soil. Eventually, SRPmust penetrate further downwards or, on a slope,laterally, before it is absorbed; in some instances, thiswill bring SRP in sub-surface flow into the surfacestreamflow. Deliberate application of excess phos-phate to augment soil fertility raises the probabilityof an ultimate augmentation of SRP exported in thedrainage water. With several studies now claimingto show that this effect is already achieved in namedcatchments (Haygarth, 1997; Smith et al., 1997),and where, in one case, the increase in bioavailabilityis substantial (Heathwaite, 1997), the risks of raisingthe fertility of receiving waters are being realised.The exports from North American wheat- andmaize-growing areas are said to include prejudicialquantities of SRP and, certainly, some of thephosphorus yields from grasslands cited by Sharpley& Rekolainen (1997; see Table 2) are perplexing intheir magnitude.

The risks will not be the same everywhere but itmay be ventured that some agricultural practicesfavour premature onset of SRP export more than doothers. For instance, tile drainage has the effect ofshortening the distance between topsoil and thepoint of interception of soil-water movement bystreams, so shortcutting its hydraulic recruitment tothe water table and shortening its exposure to deepP-binding sites. By analogy, the improvement ofgrassland on hilly, impervious catchments, carryingthin soils, potentially favours premature exhaustionof the soil binding capacity.

The possibility, signalled by Foy et al. (1995), thatSRP leakage from agriculture to surface watersthreatens to become a major agent in the eutro-phication of receiving waters, has to be consideredseriously. Its magnitude, relative to the well-known

effects of point-source urban waste waters needs tobe re-appraised. The need for catchment-specificfertiliser protocols and nutrient efflux models isbecoming increasingly clear, as a first step towardsresponsible and sustainable management of ex-ploited ecosystems.

VI. DISCUSSION

(1) Phosphorus fluxes to water and theirimpacts

The anthropogenic eutrophication of inland andcoastal waters is a pervasive, world-wide issue,manifesting itself almost everywhere that there is asignificant concentration of humanity or where thevegetation carrying-capacity of the land surface issignificantly and persistently engineered. Diagnosis,attribution, containment and correction of theconsequences are matters of national and inter-national legislation. There is a certain inevitabilitythat more people undertaking more aspiring lifestylesand making more exploitative demands on theplanet’s resources leads to accelerated biosphericdegradation and entropic dissipation of its assets.

Nevertheless, eutrophication is manageable, notby following a generic, global approach but onebased on a discipline of local diagnoses, takingaccount of local circumstances and ecological sensi-tivities and selecting locality-specific remedies. Thisentreaty is motivated primarily in respect of theliterature reviewed herein which confirms that, atevery step on its geochemical path from thelithosphere to the hydrosphere, the fate and bio-logical consequences of accelerated phosphorusfluxes are beset with alternatives : whether phos-phorus is reactive or chemically immobilised; whe-ther it is moved in solution or in particles ; whetherits transformations are subject to redox changes ; andso on. These bifurcations, or ‘ singularities ’, influencethe eventual bioavailability of phosphorus and, thus,its impacts upon the quality of aquatic environments.To be able to bias the flux pathways at each relevantbifurcation should be the vital aim of eutrophicationmanagement.

(2) The singularity of variable bioavailability

That terrestrial plants, crops, soil micro-organismsand aquatic algae all seem to have evolved veryefficient uptake mechanisms, with orders-of-mag-


nitude excess capacity over consumption, is in-dicative of the fact that BAP is naturally a scarceresource of rate- and capacity-limiting proportions.Mass-action dissociation of the oxide- and clay-bound complexes, together with the ability of someorganisms to produce alkaline phosphatases, mar-ginally enhances the BAP fraction but SRP isreasonably regarded to approximate to the spareresource-carrying capacity of phosphorus to allprimary producers.

Mostly, the resource is too precious to bothterrestrial and aquatic organisms for them to allow itto remain in free solution. Movement of SRP inmeasurable concentrations in soil water or indrainage-water (exceeding say 0.3 µM) represents abioavailability in excess of the immediate satiationrequirement of the existing biomass and of thechemical binding capacity to which it is exposed. Inopen water, however, where chemical binding sitesare relatively diluted, SRP in excess of biologicaldemand is likely to persist and the potential fertilityit confers is transported on in the gravitational flowfrom the water body to downstream water bodies.

By definition (Gibson, 1971), enriching the phos-phorus supply to a phosphorus-limited system willraise its productivity and biomass-supportive po-tential proportionately, at least to the capacity of thenext most scarce resource. This is as true forterrestrial systems, where the fertilisation is de-liberate, as it is in water, where extra productivity istransiently beneficial but mostly unwelcome(Reynolds, 1992). The rider, ‘ to the capacity of thenext most scarce resource’, is emphasised, because inneither case is the effect a continuous linear functionof phosphorus dose but a ceiling at which the crop(Heckrath et al., 1997; Johnston & Poulton, 1997a,b) or the phytoplankton (Reynolds, 1992) biomassresponse becomes saturated.

In agricultural soils, chemical binding of phos-phate represents a substantial removal mechanism;it is usually a sink for fertility. Adding phosphaticfertilisers, of whatever kind, is simply a means ofaugmenting the pools of potentially soluble ortho-phosphate, in order to raise the fluxes of assimilablephosphate at the times when the demands of cropgrowth are greatest. So long as the artificiallyenhanced bioavailability continues to be convertedto croppable biomass and the balance continues tobe inactivated through chemical sorption, theoryprovides that the bioavailable forms of orthophos-phate remain scarce in the soil water itself. It is notinevitable that the application of agricultural ferti-lisers contributes to the eutrophication of water

supplies, so long as the capacity of the crop and thesoil chemistry is not exceeded.

(3) The singularity of P transfer from landto water

Phosphorus retentivity of terrestrial environments issubvented, either grossly or insidiously, throughseveral identifiable pathways. During the last cen-tury of increasing urbanisation and sanitation, themost significant of these has been the increasedtreatment of metabolic wastes. It is accepted that themost significant of these, by far, have arisen throughthe concentration of human populations and theiractivities. Whereas the metabolic wastes of all othersignificant terrestrial animal species, living at popu-lation densities within the capacity of the localresource base, are processed as an integral part of thesame terrestrial system, the organisation of humancommunities, with attendant focusing of agriculturalfoodstuffs, necessitated the institutionalisation ofdomestic sewage disposal. The implementation ofsecondary disinfection treatments resulted in theactivation of a highly effective alternative pathwaytransferring terrestrially generated P as SRP ca-pacity to water.

As the tertiary, phosphate-removing treatment ofurban sewage becomes increasingly prevalent, therenewed interest in direct land-to-water transfersfocuses on the enhanced TP fluxes. So far as can bejudged on the evidence reviewed here, they continueto be dominated by the transport of PP. Manypotential contributions are identified, includingvegetal litter, manure and slurry and (at times,significantly) eroded soil particles. Arable fields andovergrazed (‘poached’) pasture yield more erodedsoil than permanent grass. The SRP yield fromparticulates is variable but probably relatively small.However, it is the ease of transfer of this smallproportion of the TP export that can become amaterial contribution to the eutrophication problem.The magnitude of the export is clearly influenced bysuch factors as rainfall intensity, soil porosity andgradient, as well as by the skill of agronomicmanagement (see Fig. 3). Interception of gravitatingsoil water by shallow impervious layers on hillsidesor by tile drainage and its emergence in surface flowsis a recognised avenue for the transport of non-sorbed SRP. A sharply rising relation between theexportable SRP and the TP in drainage water is asensitive indicator of a failure of soil sorptive capacity(Heckrath et al., 1995; Sibbesen & Sharpley, 1997)and should demand immediate corrective steps.


excessive application of inorganic fertiliser, manure or compost

overgrazing of improved grass on hillsides

ploughing downslope

waterlogging (increases surface flow)

compaction (increases surface flow)

drying (promotes loss of structure and eases entrainability,

(a) Surface PP loss

INCREASED RISK

excessive development of topsoil SRP pool

waterlogging (reducing conditions with P liberation)

shallow interception (title drainage, high water table, thin soil)

(b) Suburface leaching

INCREASED RISK

fertiliser applications balanced against crop uptake

level ground

DECREASED RISK

DECREASED RISK

fertiliser applications balanced against crop uptake

soil has high organic-matter content

soil has high Al-, Fe- or clay-mineral status

Fig. 3. Factors influencing the risk of transfer of phosphorus from soil to water (a) in surface run-off and (b) insubsurface leachates.

(4) The singularity of aquatic phosphorusprocessing

The conversion of delivered phosphorus to aquaticbiomass constitutes the third main singularity.Although the most familiar guidelines (OECD,1982) relate quality to the TP content, it is clearthat, as elsewhere, the fractionation is crucial to thebioavailability of P. If the objective is to maintainhigh-quality, high-clarity oligotrophic water, un-sullied by particulates or by algal turbidity exceeding2–3 mg chlorophyll a m−$, BAP concentrations haveconsistently to be below 1 mg P m−$. Many of themost oligotrophic boreal lakes satisfy this conditiondespite having rather higher average TP levels(% 10 mg P m−$). The point is that particle-boundphosphorus contributing most of the TP loads tothese lakes plays a small part in their biologicalproductivity : the plankton biomass remains strongly

constrained by the BAP. In common (but erroneous)parlance, the ‘ lakes are P-limited’.

It takes little more than an order of magnitudedifference in mean BAP to achieve the OECD(1982) guideline concentration range for eutrophiclakes (35–100 mg TP m−$). The concentrations ofnon-available phosphorus are not necessarily largerthan in oligotrophic lakes but they may represent asmaller proportion of TP. The increment in TPneeded to turn an oligotrophic lake eutrophic has tobe bioavailable (Sas, 1989), that is, supplied mainlyas readily available SRP. Given that algal concen-trations greater than the equivalent of approxi-mately 10 mg chlorophyll m−$ can be consideredtroublesome on amenity and water-treatmentgrounds, that any amount of potentially toxic,bloom-forming cyanobacteria is considered un-desirable and that the threshold of their occurrencein temperate lakes may be lower than 15 mg TP m−$


(probably lower in tropical lakes), there are goodgrounds for regarding any loads sustaining BAPconcentrations of this order as being prejudicial toaquatic ecosystem health. It is as well to bear inmind that the onset of problems of eutrophicationcan develop at BAP concentrations two to threetimes smaller (5–10 mg BAP m−$). Turning therelationship around, a phytoplankton crop equiva-lent to 10 mg chlorophyll m−$ is sustainable on aBAP concentration of approximately 2.5 mg m−$.For a catchment yielding a net run-off equivalent of500 mm year−", the removal of an average 1.25 mgSRP m−# is sufficient to raise the concentration of therun-off to threaten diminished water quality inreceiving systems.

The mean biomass of the phytoplankton istypically less than the maximum supportive capacityof the annual BAP flux, especially if algal mass isconverted to consumer or decomposer mass, or isrecruited to the sediment in particulate form. Thus,the longer the water is retained in a lake, the moreof the original BAP load is likely to be processed intobiogenic sedimentary fractions or sorbed onto de-posited oxides and hydroxides. Sedimentation is,ordinarily, a P-removing process, capable of dilutingthe risk of deleterious impacts of the phosphorusexported from the catchment. Were this not so, theVollenweider model could not hold. As argued inSection IV, the efficiency of this immobilisation is,by analogy with the events in catchment soils,influenced by the rate of mineralisation and by theavailability of chemical binding sites in the super-ficial sediments.

The magnitude of phosphorus re-use in aquaticsystems and of the subsequent yields of biomassvaries with the local conditions for recycling and theeffectiveness of each of the possible contributorymechanisms (see Section IV.7, above). These con-ditions are also tabulated in summary form andlisted in terms of risk to water quality in Fig. 4.

Thus, with low phosphorus inputs, the biologicalresponses of receiving waters comply well with theVollenweider-type long-range average behaviour.Up to a point, enhanced loadings from anthro-pogenic sources induce broadly predictable re-sponses. Within this range, the extent of loadcorrections are also conveniently indicated. How-ever, Sas’ (1989) analyses of case studies reminds usthat the responses may be, at best, delayed or, atworst, practically unattainable. The decisive factoris how easily the sedimenting phosphorus is returnedto productive consumption and how often thisbioavailability is renewed. The most sensitive systems

in this respect are those with the least sedimentaryfluxes of binding capacity relative to those oforganismic phosphorus and the most frequent oppor-tunities to return BAP to the water column.Shallowness, frequency of turbation and, especially,failure of sorption capacity each contribute to therecyclability of phosphorus in lakes and rivers.

(5) Yield equivalence of phosphorus sources

Perhaps the most useful perspective to developconcerns the cumulative effects of these singularitiesin relating aquatic sensitivity to phosphorus source.Many factors contribute to the difficulties of doing soand to our trepidation about offering a firstquantitative estimate. The caveats are numerous inas much as the outcome should take account ofgeology, geomorphology, and the history of land useand fertiliser application. The yields of phosphorusfrom catchment to water depend on: the amountand type of fertiliser applied, the timing of itsapplication, the nature of the soil, the depth of thesoil, the slope of the ground, the extent of the plantcover, the intensity and aggregate of rainfall, theevaporative residual and the proximal fate of thedrainage water. Once liberated into a discrete bodyof water, the extent to which the phosphoruspromotes undesirable plant growth depends on theproximal source of the phosphorus and whether it isin solution, readily soluble or so tightly bound toparticulates that it has no novel effect at all. If thephosphorus is bioavailable in the receiving water,the undesirability of its effects is subject to suchvariables as hydraulic retention and the opportunityfor it to be invested in either microscopic ormacrophytic plant mass. The fate of this now-organismic phosphorus itself depends on water depthand the amenability of the material to decompositionbefore it has been buried in the bottom sediment bythe settlement of non-reactive particulates or beforesome of the resultant soluble phosphorus is dis-charged again into the trophogenic zone of biologicalre-use.

An alternative approach is to start with whatis, stoichiometrically, the best authenticatedof the relationships. We take, as an arbitrarybase, the eutrophy-threshold BAP concentration(35 mg P m−$ : Foy & Withers, 1995, based onOECD, 1982), at which the standing stock ofphytoplankton arguably reaches nuisance pro-portions. In order to relate this concentration to theremoval rate in drainage water, we have to nominatea catchment run-off (precipitation less evapotrans-


sediments shallow, subject to wave action or entrainment

sediments fine, calcareous, poor in iron

water overstocked with benthivorous fish

benthic reducing conditions

HIGH RISK

well-balanced fish fauna

iron-rich, low-organic carbon sediments, with low BOD

deep oxic sediments, beyond entrainment

LOW RISK

Fig. 4. Factors influencing the risk of transfer of phosphorus recycling from previously sedimented phosphorus back tothe water column.

piration: say, 0.5 m year−" ; we suppose this to beappropriate to north-west Europe). To support amean concentration of 35 mg P m−$, the hydraulicyield, 0.5 m$ m−# year−", must elute a 17.5 mgSRP m−# year−". In drier climates, the sustainingelution is proportionately less. Comparison with theSRP losses claimed for the series of land uses noted inTable 2 indicates that source waters delivered fromcertain kinds of grassland and from many well-established arable fields are, if extrapolated acrossentire lake catchments, fully capable of achieving thenominated threshold of eutrophy in receiving waters.

Against this, we recall from Section V.2 thatapplications of inorganic P fertiliser to unsaturatedsoils (in the range 2–10 g m−# year−") and cropremoval (equivalent to 1–3 g P m−# year−"), repu-tedly yield trivial quantities of SRP to drainage,generally less than 1% of the TP applied. Thebalance is supposed to be retained at chemicalbinding sites in the soil and subsoil. We have arguedthat any such SRP leakage is prejudicial to waterquality and concern should focus on the concen-tration of SRP in surface run-off that is not exposedto binding capacity, the SRP in sub-surface run-offthat is not inactivated by the current bindingcapacity and the extent to which the accumulationof P bound in soil anticipates eventual saturation ofthe binding capacity (the ‘change point ’ of Heckrathet al., 1995, 1997), beyond which substantial SRPleakage to drainage will occur.

While the largest terrestrial transfers to waterinvolve phosphorus fractions other than SRP, prob-lems with its bioavailability are largely unrealised.We deduce that the biological effect of terrestrial TP

is mainly confined to the supposed 1% SRP yield. Atface value, then, to generate a concentration indrainage water of 35 mg P m−$ requires a TP exportof 1750 mg m−#. The entries in Table 2 show that TPlosses of this magnitude are shed from someintensively arable croplands.

Comparison of these figures with the yield fromsecondary treatment invokes the approximation ofMorse et al. (1993) for the per capita humangeneration of phosphorus (0.58 kg P ind−" year−",excluding any other detergent or waste P yields) andthe typical 70–100% BAP clearance of the treat-ment plant. The yield of 17.5 mg BAP m−#

(17.5 kg P km−#) is thus equivalent to a residentpopulation density of 30–44 persons km−#.

As is by now well-recognised, exceedence of theeutrophication threshold in the secondary-treatedwaste water is difficult to avoid, especially whensome urban populations occupy land at densities ofover 10000 persons km−# and are thus responsible forannual outputs of 4–6 tonnes of SRP per squarekilometre. Implementation of tertiary sewage treat-ment, in order to secure a much lower phosphoruscontent in urban waste waters (actual efficienciesvary but 85–90% is generally achievable), alters thecalculated population equivalences based on eutro-phic thresholds by a factor of 7–10, to over 200persons km−#.

(6) Epilogue

Acknowledging the difficulties and dangers of com-parisons among contributory sources, the calcula-tions in the preceding section provide some per-


spective on the relative magnitude of threats towater quality posed by urban growth and agri-cultural intensification. On the scale of humandevelopment since the last glaciation, the clearanceof forest, the imposition of agriculture and the stepstaken to improve the fertility and agricultural yieldsfrom soils have, together, promoted the flux ofnutrients to aquatic systems and stimulated theirbiological productivity. This is still small in com-parison with the widespread eutrophication of themost recent quinquennium which is attributableprimarily to the widespread adoption of secondarytreatment of sewage generated by a rapidly ex-panding, increasingly urbanised human population.Part of that expansion has been sustained by agreatly improved capacity for food production, itselffurnished by the supplementation in abundance ofsoil nutrients with natural sources of nitrate andphosphate. As with several other technical achieve-ments of our species, they involve a scale of planetaryengineering with uncontemplated consequences.

So far as lake catchments are concerned, there arealready changes which redress the balance. Tertiarytreatment will become more widespread and will besupplemented by new techniques of waste-waterrecovery while the capacity of soils to go on sorbingphosphorus is increasingly challenged. Policies forthe fertility of food-producing land, almost on afield-to-field basis need to be developed to ensurethat fertiliser applications match the demand ofoptimal crop yield without exceedence of the soil’scapacity to avert the risk of leaching of bioavailablephosphorus. If it is considered important that thesupportive capacities of fresh waters are containedand that a stage shift to chaotic phosphorus transfersto drainage waters is avoided, then it is imperativethat new understanding about the impacts ofglobally enhanced nutrient fluxes is embodied intooverall environmental management policies.

VI. CONCLUSIONS

1. The role of phosphorus as a primary agent ineutrophication is confirmed. The capacity to supportadditional standing biomass of water plants or,especially, of phytoplankton is linked directly to theanthropogenic acceleration of the rates of loadingbiologically available phosphorus fractions in drain-age waters.

2. Biological availability (or bioavailabilty) ofphosphorus refers, in essence, to those fractions of thetotal mass of phosphorus present in a system that arereadily assimilable by organisms, or are made more

assimilable through the activities of the organismsthemselves (for instance, through the production ofphosphatases), and that portion of the phosphoruswhich has been assimilated and is already intra-cellular. Analytical determinations of soluble, molyb-date-reactive phosphorus (SRP) usually underesti-mate the amount of immediately bioavailable P(BAP). However, the preferred measurements oftotal phosphorus (TP) include substantial fractionsof scarcely soluble phosphates of alkaline earthmetals, phosphates bound to metal oxides andhydroxides and in organic residues which arescarcely or only conditionally available to algae andplants. Thus TP measurements tend to exaggerateBAP.

3. BAP is introduced into streams, rivers andlakes, primarily as a consequence of aqueous leachingfrom the hydrological catchment. The quantities ofusable phosphorus available are related firstly tothe geochemistry of the catchment and the mineraldeposits present, then to rates of their erosion, whichwill be influenced by geomorphology, climate andhydrology. The extent of sorption of liberatedphosphate by catchment soils and of its assimilationinto terrestrial biomass will further influence theamounts transferred to water.

4. The phosphorus reaching streams, rivers andlakes is, almost everywhere, subject to enrichmentfrom anthropogenic sources. The major en-hancements in SRP levels are attributable primarilyto solubilisation of phosphorus in secondary sewagetreatment. Levels are especially high in riversdraining urban catchments. Efficient tertiary treat-ment of sewage effluents ‘phosphorous stripping’can remove up to 95% of the SRP. TP outputs fromagricultural land are also high and, in some areas,are rising. The predominant fraction is often at-tributable to eroded soil particles which offer littleBAP. Losses of SRP in surface run-off (overlandflow) are enhanced by some agricultural practices,and influenced by slope and drainage; loss insubsurface leaching is influenced by the depth andmineral content of the soil. The P-binding capacityof soils varies, is theoretically saturable and bindingsites can be by-passed by free or inappropriatedrainage works. Measures to contain potentiallylarge BAP removal from soil to water are alreadynecessary in some localities.

5. The direct effects of BAP on deeper standingwaters are generally well summarised by the Vollen-weider-OECD model. This is because the fate ofmost of the phosphorus reaching deep sediments is tobecome chemically bound, mainly with iron hy-


droxide, and to be finally buried there. Redoxchanges can result in re-solution of the phosphorusbut most of this reprecipitates again without stimu-lating significant new biological production. Themodel does not predict biomass well in lakes wherephosphorus is not the primary limitation on bio-logical carrying capacity, which include thosealready enriched with phosphorus to the extent thatthe next limiting factor is reached. The model maynot predict well the responses to additional phos-phorus loads in low-latitude lakes or in shallow lakeswhere unbound phosphorus is physically recycledbetween sediments and lake water. The Vollen-weider model is not a good basis for phosphorusmanagement in any individual lake.6. No general model of whether a lake sedimentbehaves as a net source of bioavailable phosphorusor as a net sink is available. Critical quantities relateto the partitioning of the external load among theoutflow and the sedimentary flux. The capacity of asediment to retain the sedimentary flux and itspredisposition to be able to return (‘ to release ’)phosphorus to the water are estimable. Saturation ofa site-specific threshold value, [P]

k,facilitates free

exchange of SRP with the water column, where itmay be recycled in the assembly of new biomass.This can continue long after the external load is cut.7. The depth of ‘active sediment’ which is involvedin exchanges probably does not exceed 40–50 mm.8. The control and management of eutrophicationin specific water bodies may be approached from thestandpoint of critical singularities and the regulationthat it is possible to bring to bear on each. Theseinclude the transfer of phosphorus from land towater and the ability of the water body to sequesterits external loads to the permanent sediments.Bioavailability relates mostly to what remains chemi-cally unbound; if there is SRP in abundance, thebiological system does not need it and it is insensitive(at least initially) to corrective treatment of the Pload. The means of reducing BAP mainly requiresite- or catchment-specific actions aimed at thecritical singularities.

VIII. ACKNOWLEDGEMENTS

This review is based on a report commissioned by theEuropean Fertiliser Manufacturers’ Association. We aregrateful for the Association’s agreement and encour-agement to publish the review in the scientific literature.PSD’s work was supported by a grant from Tesco Ltd.The paper benefitted from the incisive comments and

suggestions of two anonymous reviews. The assistance ofthe Editor is also gratefully acknowledged.

IX. REFERENCES

A, T. (1975). Effects of man-induced and natural loadings of

phosphorus and nitrogen on the large Swedish lakes.

Verhandlungen der internationale Vereingung fuX r theoretische und

angewandte Limnologie 19, 1125–1132.

A, T. (1988). Background yield of phosphorus from drainage

area and atmosphere. Hydrobiologia 170, 35–44.

A, I. (1977). Role of sediments in the process of recovery

of a eutrophicated lake. In Interactions between sediments and

freshwater (ed. H. L. Golterman), pp. 372–377. Junk, Den

Haag.

A, S. E., C, A., W, E. J. & E, C. C.

(1968). The plant-nutrient content of rainwater. Journal of

Ecology 56, 497–504.

B, D. S. (1996). The phosphorus composition of a

diverse series of Australian sediments. Hydrobiologia 335,

63–73.

B, D. S., B, J. K., C, L. M. & J,

D. R. (1995). Phosphate ester hydrolysis facilitated by mineral

phases. Environmental Science and Technology 29, 1706–1709.

B, J. & R, D. A. (1985). Soil erosion, climatic

vagary and agricultural change on the Downs around Lewes

and Brighton. Applied Geography 5, 234–258.

B$ , B. & P, K. (1982). Different patterns of

phosphorus release from lake sediments in laboratory experi-

ments. Hydrobiologia 92, 415–429.

B, N. C. (1990). The nature and properties of soils (Tenth

edition). Macmillan, New York.

B, P. C., P, D. S. & J, D. S. (1984).

Phosphorus in the soil microbial biomass. Soil Biology and

Biochemistry 16, 169–175.

B, P. H. & L, C. (1997). The flow of nutrients in

the Danube River basin. EAWAG News 43, 15–17.

B, D. (1979). The continuous culture of phytoplank-

ton: mathematical equivalence among three steady-state

models. American Naturalist 113, 123–134.

B, D. K. (1985). Kinetics of nutrient-limited transport

and microbial growth. Microbiological Reviews 49, 270–297.

C, D. E. & B, R. W. (1981). Prediction of total

phosphorus concentration, chlorophyll a and Secchi depth in

natural and artificial lakes. Canadian Journal of Fisheries and

aquatic Science 38, 414–423.

C, N. F. (1995). Influence of human populations on P

transfers to aquatic systems: a regionl scale study using large

rivers. In Phosphorus in the global environment (ed. H. Tiessen),

pp. 235–244. John Wiley & Sons, Chichester.

C, S. R. (1980). Enrichment of Lake Wingra, Wis-

consin, by submersed macrophyte decay. Ecology 61, 1145–

1155.

C, S. R. (1981). Submersed vegetation: an internal

factor in lake ecosystem succession. American Naturalist 118,

372–383.

C, A. D., A, N. J. & H, P. J. (1984). The

utilisation of inorganic phosphorus compounds as nutrients by

eukaryotic microalgae: A multidisciplinary perspective, Criti-

cal Reviews in Microbiology 10, 317–391.


C, S. C. & J, M. L. (1957). Fractionation of soil

phosphorus. Soil Science 84, 133–144.

C, G. & S, P. (1984). Nutrient losses by leaching

and runoff and possibilities for their control. In Nutrient

balances and fertilizer needs in temperate agriculture (ed. Anon.), pp.

137–156. International Potash Institute, Basel.

C, N. L. & L, G. F. (1965). Hydrolysis of condensed

polyphosphates. I. Non-sterile environment. Air and Water

Pollution 9, 723–742.

C, G. W. (1976). A review of the the effects of agriculture

on the chemical composition and quality of surface and

underground waters. Technical Bulletin, Ministry of Agriculture,

Fisheries and Food 32, 5–57.

C, G. W. & W, R. J. B. (1973). Significance of

man-made sources of phosphorus : fertilisers and farming.

Water Research 7, 19–33.

C, P. A., R, C. S., L, A. J. & L,

J. P. (1994). Sediment analysis of sensitive Cumbrian lakes.

Unpublished report to The National Rivers Authority.

Institute of Freshwater Ecology, Ambleside.

C, C. & S, G. (1997). Binding of DNA on humic

acids : effects on transformations of Bacillus subtilis. Soil

Biology and Biochemistry 30, 1061–1067.

D, P. S. (1997). Fluxes of phosphorus relevant to phytoplankton

ecology. Doctoral thesis, University of Leeds.

D, L. A. (1938). An attempted fractionation of the soil

phosphorus. Journal of agricultural Science 28, 234–246.

D, P. J. & K, W. B. (1974). The effects of geology

and land-use on the export of phosphorus from watersheds.

Water Research 9, 135–148.

D, P. J. & R, F. H. (1974). The phosphorus-

chlorophyll relationship in lakes. Limnology and Oceanography

19, 767–773.

D, E. , G, R. D., I, A. & M, J. T.

(1993). Macrophyte-related shifts in the nitrogen and phos-

phorus contents of the different trophic levels in a biomanipu-

lated shallow lake. Hydrobiologia 246, 19–26.

D, J. C. & H, S. I. (1987). Occurrence of phosphorus

and its potential mobilisation in the littoral sediments of a

productive English lake. Freshwater Biology 17, 513–523.

D, M. R. (1973). Some thoughts on nutrient limitation in

algae. Journal of Phycology 9, 264–272.

D, M. R. (1974). The nutrient status of algae grown in

continuous culture. Journal of the Marine Biological Association of

the United Kingdom 54, 825–855.

D, R. C. (1967). Nutrient limitation in the sea:

dynamics, identification and significance. Limnology and

Oceanography 12, 685–695.

E, W. (1936). U$ ber die Beziehungen des Eisenkreislaufs

zum Phosphatkreislauf im eutrophen See. Archiv fuX r Hydrobio-

logie 33, 361–387.

E, P. (1994). Bioavailability of phosphorus in agri-

culturally loaded rivers in southern Finland. Hydrobiologia

287, 179–194.

E P. (1998). Algal-available phosphorus originating from

agriculture and municipalities. Monographs of the Boreal

Environment Research 11, 7–60.

E, J. (1980). The phosphorus cycle. In The handbook of

environmental chemistry, Volume 1A (ed. O. Hutzinger), pp.

147–167. Springer Verlag, Berlin.

E, D. R. & W, M. E. (1984). Chemical stra-

tigraphy of lake sediments as a record of environmental

change. In Lake sediments and environmental history (ed. E. Y.

Haworth and J. W. G. Lund), pp. 11–67. Lancaster Uni-

versity Press, Lancaster.

E, R. & N, S. (1978). Soil erosion in north

Norfolk. Journal of Agricultural Science 90, 185–192.

E, J. J. (1983). Succession. In Tropical rain forest ecosystems

(ed. F. B. Golley), pp. 217–223. Elsevier, Amsterdam.

F, G., F, R. & S, A. J. (1989). Bio-

energetic characterisation of transient state phosphate uptake

by the Cyanobacterium Anacystis nidulans. Theoretical and

experimental basis for a sensory mechanism adapting to

varying environmental phosphate levels. Archives of Micro-

biology 152, 353–361.

F, B. (1996). Pattern and process in neotropical rain

forests : the first 100 years of succession. Trends in Ecology and

Evolution 11, 119–124.

F, T. R. & L, D. R. S. (1992). Interpretation of

radiophosphate dynamics in lake water. Canadian Journal of

Fisheries and aquatic Sciences 49, 252–258.

F, R. H. & L, S. D. (2000). Contributions of diffuse

and point sources to phosphorus loads in the River Main over

a 22-year period. Boreal Environmental Research 5, 27–37.

F, R. H., S, R. V., J, C. & L, S. D. (1995).

Upward trend in soluble phosphorus loadings to Lough

Neagh despite phosphorus reduction at sewage treatment

works. Water Research 29, 1051–1063.

F, R. H. & W, P. J. A. (1995). The contribution of

agricultural phosphorus to eutrophication. Proceedings of the

Fertiliser Society 365, 1–32.

F, D. A. (1990). Alteration of bioavailability and toxicity

by phototransformation of organic acids. In Organic acids in

aquatic ecosystems (ed. E. M. Perdue and E. T. Gjessing),

pp. 167–177. John Wiley & Sons, Chichester.

G, C. E. (1971). Nutrient limitation. Journal of the Water

Pollution Control Federation 43, 2436–2440.

G, C. E. (1997). The dynamics of phosphorus in freshwater

and marine environments. In Phosphorus loss from soil to water

(ed. H. Tunney, O. T. Carton, P. C. Brookes and A. E.

Johnston), pp. 119–135. CAB International, Wallingford.

G, C. E. & W, Y. (1995). Substance budgets of an upland

catchment: the significance of atmospheric phosphorus inputs.

Freshwater Biology 33, 385–392.

G, H. L., B, C. C. & J-M$ , J. (1969).

Availability of mud phosphates for the growth of algae.

Verhandlungen der internationale Vereinigung fuX r theoretische und


G, J. U. & H, W. A. (1995). Phosphorus as a

limiting resource in inland waters ; interactions with nitrogen.

In: Phosphorus in the global environment (ed. H. Tiessen), pp.

255– 273. John Wiley & Sons, Chichester.

H, D. A. & S, L. L. (1987). Generalised water-

shed loading functions for stream flow nutrients. Water

Resources Bulletin 23, 471–478.

H, D. P. & M, S. F. (1997). Wave-induced

shear stresses, plant nutrients and chlorophyll in seven shallow

lakes. Freshwater Biology 38, 159–168.

H, D. M. (1992). Eutrophication of freshwaters. Chapman &

Hall, London.

H, P. M. (1997). Agriculture as a source of phosphorus

transfer to water : sources and pathways. Scope Newsletter 21,

1–16.

H, J., F, R. H. & G, C. E. (1993). Nitrogen and

phosphorus budgets in the Erne system, 1974–1989. Proceedings

of the Royal Irish Academy 93B, 33–44.


H, A. L. (1997). Sources and pathways of phos-

phorus loss from agriculture. In Phosphorus loss from soil to water



H, G., B, P. C., P, P. R. & G,

K. W. T. (1995). Phosphorus leaching from soils containing

different P concentrations in the Broadbalk experiment.

Journal of environmental Quality 24, 904–910.

H, G., B, P. C., P, P. R. & G,

K. W. T. (1997). Phosphorus losses in drainage water from an

arable silty clay loam soil. In Phosphorus loss from soil to water



H, A. H. M. & L, L. (1980). Fractionation of

inorganic phosphate in calcareous sediments. Journal of

Environmental Quality, 9, 405–407.

H, J. E. & L, G. E. (1973). Output of phosphorus,

dissolved organic carbon and fine particulate carbon from

Hubbard Brook watersheds. Limnology and Oceanography 18,

734–742.

H, H., K-N, L. & S, A. O. (1988).

Phosphorus in soil, water and sediment. Hydrobiologia 170,

19–34.

H, W. A. (1990). The prediction of phosphate coprecipi-

tation with calcite in freshwaters. Water Research 24, 1017–

1023.

H, R. W., J, H. S., M, R. & P, H.

(1995). Transport to and processing of P in near-shore and

oceanic waters. In Phosphorus in the global environment (ed. H.

Tiessen), pp. 323–345. John Wiley & Sons, Chichester.

H, T. (1994). Suspended sediment transport in Lake

Sa$ kyla$ n Pyha$ ja$ rvi. Aqua Fennica 24, 171–185.

J, W. F. & B, J. W. (1991). Littoral-pelagic phos-

phorus dynamics during night-time convective circulation.

Limnology and Oceanography 36, 949–960.

J, W. F. & B, J. W. (1997). Net and gross sedi-

mentation in relation to the phosphorus budget of Eau Galle

Reservoir, Wisconsin. Hydrobiologia 345, 15–20.

J, J. P., K, P., J, E. & S, A.

(1992). Iron-phosphorus ratio in surface sediment as an

indicator of phosphate release from aerobic sediments in

shallow lakes. Hydrobiologia 235/236, 731–743.

J, W. J. (1994). Resource-recovery wastewater treatment.

American Scientist 82, 366–375.

J, W. J. & MC, P. L. (1971). Aerobic decompo-

sition of algae. Environmental Science and Technology 5,

1023–1031.

J, A. E. & P, P. R. (1992). The role of

phosphorus in crop production and soil fertility : 150 years of

field experiments at Rothamsted, United Kingdom. In

Proceedings of the International Workshop on Phosphate Fertilizer and

the Environment (ed. J. J. Schultz), pp. 45–63. IFDC, Tampa.

J, A. E. & P, P. R. (1997a). Defining critical

levels of available soil phosphorus for agricultural crops. In

Phosphorus loss from soil to water (ed. H. Tunney, O. T. Carton,

P. C. Brookes and A. E. Johnston), pp. 441–445. CAB In-

ternational, Wallingford.

J, A. E. & P, P. R. (1997b). Factors affecting

critical soil phosphorus values. In Phosphorus loss from soil to

water (ed. H. Tunney, O. T. Carton, P. C. Brookes and A. E.


J, W. R., I, F., D, R. M. & P,

A. F. (1965). Nitrogen and phosphorus in tile drainage

effluent. Proceedings of the Soil Science Society of America 29,

287–289.

J, R. I., S, P. J. & D H, H. (1993). Effects of

dissolved humic substances on the speciation of iron and

phosphate at different pH and ionic strength. Environmental

Science and Technology 27, 1052–1059.

K, K. (1994). Effect of summer weather on internal

loading and chlorophyll a in a shallow lake: a modelling

approach. Hydrobiologia 275}276, 371–378.

K$ , T. & B, D. (1990). Biological availability

of phosphorus in agricultural runoff compared to other

phosphorus sources. Verhandlugen der internationale Vereinungen

fuX r theoretische und angewandte Limnologie 24, 214–217.

K, P. & S, A. J. C. (1992). Phosphorus in the sediment

of the Loosdrecht lakes and its implication for lake restoration

perspectives. Hydrobiologia 233, 39–50.

K, J. V., E, D. N., S, P. E. & R,

D. M. (1997). Sedimentary phosphorus cycling and a phos-

phorus mass balance for the Green Bay (Lake Michigan)

ecosystem. Canadian Journal of Fisheries and Aquatic Sciences 54,

10–26.

K, G. J. (1972). The eutrophication of surface

water by agriculture and the urban population. Stickstoff 15,

56–67.

L, V. A. (1975). Digestive activities of carp as a major

contributor to nutrient loading of lakes. Verhandlungen der

internationale Vereingung fuX r theoretische und angewandte Limnologie

19, 2461–2468.

L, D. R. S. (1973). Movements of phosphorus between its

biologically important forms in lake water. Journal of the

Fisheries Research Board of Canada 30, 1525–1536.

L, S. D., F, R. H., S, R. V. & J, C. (1997).

Estimating the contribution from agriculture to the phos-

phorus load in surface water. In Phosphorus loss from soil to water


Johnston), p. 55–75. CAB International, Wallingford.

L, L. (1977). The role of iron in the exchange of

phosphate between water and sediments. In Interactions between

sediments and freshwater (ed. H. L. Golterman), pp. 313–317.

W. Junk, Den Haag.

L, J. W. G. (1965). The ecology of the freshwater phyto-

plankton. Biological Reviews of the Cambridge Philosophical Society

40, 231–293.

L, J. W. G. (1972). Eutrophication. Proceedings of the Royal

Society, London B 180, 371–382.

M, F. J. H. (1963). Some methods of water analysis

for limnologists. Scientific Publications of the Freshwater Biological

Association No. 21: 71 pp.

M, J. C. & B, P. (1991). Evidence for the

restoration of the Lake Erie ecosystem. BioScience 41, 216–223.

MC, E., D, J. A. & W, S. (1989). Sig-

moidal relationships between nutrients and chlorophyll

among lakes. Canadian Journal of Fisheries and aquatic Sciences 46,

1171–1175.

MC, E. & K, J. (1981). Empirical relationships

between phytoplankton and zooplankton biomass in lakes.

Canadian Journal of Fisheries and aquatic Sciences 38 : 458–463.

M, N. H. (1995). How do cells express nutrient limitation at

the molecular level? In: Molecular ecology of aquatic microbes (ed.

I. Joint), pp. 171–190. Springer-Verlag, Berlin.

M, M. L., J, E., D, E. & M, B. (1994).

Long-term responses to fish-stock reduction in small, shallow


lakes – interpretation of five-year results of four bio-

manipulation cases in the Netherlands and Denmark. Hydro-

biologia 276, 457–466.

M, E. L., & S, A. (1982). Evaluation of fish

communities through assessment of zooplankton populations

and measures of lake productivity. North American Journal of

Fisheries Management 2, 14–27.

M, D. T. & B, P. C. M. (1994). Influence of

internal loading on phosphorus concentration in shallow lakes

before and after reduction in external loading. Hydrobiologia

275/276, 379–389.

M, D. T. , B, A., B, P. C. M. &

R, C. W. J. (1997). Dutch policy towards phosphorus

losses in agriculture. In Phosphorus loss from soil to water (ed. H.

Tunney, O. T. Carton, P. C. Brookes and A. E. Johnston),

pp. 407–409. CAB International, Wallingford.

M, M. A. (1997). The behaviour of soil and fertiliser

phosphorus. In Phosphorus loss from soil to water (ed. H. Tunney,

O. T. Carton, P. C. Brookes and A. E. Johnston), pp. 137–

149. CAB International, Wallingford.

M, M. F. (1941). Chemical soil diagnosis by the universal

soil testing system. Bulletin of the Connecticut agricultural

experimental Station, No. 450, 35pp.

M, G. K., L, J. N. & P, R. (1993). The economic

and environmental impact of phosphorus removal from wastewater in the

European Community. Selper Publications, London.

M, C. H. (1941). The exchange of dissolved substances

between mud and water in lakes. I. Journal of Ecology 29,

280–329.

M, C. H. (1942). The exchange of dissolved substances

between mud and water in lakes. II. Journal of Ecology 30,

147–201.

M, J. & R, J. P. (1962). A modified single-solution

method for the determination of phosphate in natural waters.

Analytica Chimica Acta 27, 1–36.

M, A. D. & W, P. J. A. (1996). The phosphorus

content of rivers in England and Wales. Water Research 30,

1258–1268.

N, S. W. (1988). Physical energy inputs and the com-

parative ecology of lake and marine ecosystems. Limnology and

Oceanography 33, 1005–1025.

N, P. & T, P. B. (1977). Solute movement in the soil-root

system. Blackwell Scientific Publications, Oxford.

O, S. R., C, C. V., W, F. S. & D, L. A.

(1954). Estimation of available phosphorus in soils by extraction with

sodium bicarbonate. U.S. Department of Agriculture, Washing-

ton.

O, J. M. (1977). Non-point source – stream nutrient level

relationships: a nationwide study. U.S. Environmental Protection

Agency, Corvallis.

O E C- D

(1982). Eutrophication of waters: monitoring and assessment.

OECD, Paris.

P, R. H. & C, A. (1984). The effects of turbulence

on phosphorus supply in a shallow bay of Lake Memphre!ma-

gog. Verhandlungen der internationale Vereinigung fuX r theoretische und


P$ , O.-P. (1997). Agricultural phosphorus load and

phosphorus as a limiting factor for algal growth in Finnish

lakes and rivers. In Phosphorus loss from soil to water (ed. H.

Tunney, O. T. Carton, P. C. Brookes and A. E. Johnston),


P$ , O.-P. & R, S. (1991). Dissolved reactive

and total phosphorus load from forest and agricultural areas

in Finland. Aqua Fennica 21, 127–136.

P, B. & D, J. M. (1997). Movement of phosphorus

from agricultural soil to water. In Phosphorus loss from soil to



P, Y. T., D, C.M & K, J. (1989). Unifying

nutrient-chlorophyll relationships in lakes. Canadian Journal of

Fisheries and aquatic Sciences 46, 1176–1182.

P, C. M. & H, D. W. (1954). Determination of

inorganic phosphate in sea water by an iso-butanol extraction

procedure. Journal of Marine Research 13, 122–132.

P, R., B$ , B., D, M., P, K., P,

R. & S, M. (1988). Fractionation of phosphorus in

suspended matter and sediment. Ergebnisse der Limnologie 30,

98–103.

R, A. C. (1958). The biological control of chemical

factors in the environment. American Scientist 46, 205–221.

R, S. (1989). Phosphorus and nitrogen load from

forest and agricultural areas in Finland. Aqua Fennica 19,

95–107.

R, C. S. (1978). Phosphorus and the eutrophication of

lakes – a personal view. In Phosphorus in the environment: its

chemistry and biochemistry (ed. R. Porter and D. Fitz-

Simons), pp. 201–228. Excerpta Medica, Amsterdam.

R, C. S. (1979). The limnology of the eutrophic meres

of the Shropshire-Cheshire Plain: a review. Field Studies 5,

93–173.

R, C. S. (1984). The ecology of freshwater phytoplankton.

Cambridge University Press, Cambridge.

R, C. S. (1992). Eutrophication and the manaagement

of planktonic alagae: what Vollenweider couldn’t tell us. In

Eutrophication: research and application to water supply (ed. D. W.

Sutcliffe and J. G. Jones), pp. 4–29. Freshwater Biological

Association, Ambleside.

R, C. S. (1996). Phosphorus recycling in lakes : evidence

from large limnetic enclosures for the importance of shallow

sediments. Freshwater Biology 35, 623–645.

R, C. S. (1997). Pelagic vegetation processes; a model for

ecosystem theory. Ecology Institute, Oldendorf.

R, C. S. (1998). Back from the brink: the restoration of

Windermere. Cumbria Wildlife 50, 8–14.

R, C. S. (1999). Metabolic sensitivities of lacustrine

ecosystems to anthropogenic forcing. Aquatic Sciences 61,

183–205.

R, C. S., R, S. N., M, I. F. &

M, M. (2000). The regulation of phytoplankton

poulation dynamics in the world’s largest lakes. Aquatic

Ecosystem Health and Management 3, 1–21.

R B., H, H., T, F. & B, C.

(1995). The role of mixotrophy in pelagic environments. In

Molecular ecology of aquatic microbes (ed. I. Joint), pp. 87–114.

Springer Verlag, Berlin.

R, F. H. (1968). Further observations inconsistent with

the hypothesis that the molybdenum blue method measures

orthophosphate in lake water. Limnology and Oceanography 13,

7–13.

R, E. T. & P, E. E. (1984). Role of internal

phosphorus loading in two shallow, productive lakes in

Alberta, Canada. Canadian Journal of Fisheries and aquatic

Sciences 41, 845–855.


R, M. L. (1971). Paradox of enrichment : destabiliz-

ation of exploitation ecosystems in ecological time. Science 171,

385–387.

R, K. (1992). Development of a sequential extraction

method for different forms of P in marine sediments. Limnology

and Oceanography 37, 1460–1482.

R, J. C., S, J. K. & H, R. F. (1973). Phosphorus

run-off in streams. Advances in Agronomy 25, 1–45.

S, H. (1989). Lake restoration by reduction of nutrient loading.

Akademia Verlag Richarz, Skt Augustin.

S, M. (1966). Primary production by phytoplankton

community in some Japanese lakes and its dependence upon

lake depth. Archiv fuX r Hydrobiologie 62, 1–28.

S, M. (1998). Ecology of shallow lakes. Chapman and

Hall, London.

S, D. W. (1977). Evolution of phosphorus limitation in

lakes. Science 196, 260–262.

S, O. F. & B, A. (1997). The relation

between accumulation and leaching of phosphorus : labora-

tory, field and modelling results. In Phosphorus loss from soil to



S, G. E., B, R. E., P, R. F. & S,

R. C. (1973). Phosphorus losses from eroding soils. Journal of

environmental Water Quality 2, 299–302.

S, A. N., H, M. J., S, E., H-

I, A.., H, W. A. & R, L. (1995).

Phosphorus transfers from terrestrial to aquatic ecosystems. In

Phosphorus in the global environment (ed. H. Tiessen), pp.

171–199. John Wiley & Sons, Chichester.

S, A. N. & M, R. G. (1987). The impact of soil

and fertilizer phosphorus on the environment. Advances in

Agronomy 41, 297–324.

S, A. N. & R, S. (1997). Phosphorus in

agriculture and its environmental implications. In Phosphorus

loss from soil to water (ed. H. Tunney, O. T. Carton, P. C.

Brookes and A. E. Johnston), pp. 1–53. CAB International,

Wallingford.

S, A. N., S, S. J., J, O. R., B, W. A. &

C, G. A. (1992). The transport of bioavailable

phosphorus in agricultural run-off. Journal of Environmental

Quality, 21, 30–35.

S, P. J. (1994). The effect of pH, dissolved humic substances

and ionic composition on the transfer of iron and phosphate

to particulate size fractions in epilimnetic lake water. Limnology

and Oceanography 39, 1734–1743.

S, E. & S, A. N. (1997). Setting and justifying

upper critical limits for phosphorus in soils. In Phosphorus loss

from soil to water (ed. H. Tunney, O. T. Carton, P. C. Brookes

and A. E. Johnston), pp. 151–176. CAB International,

Wallingford.

S, M. I. (1995). Signal transduction in microorganisms. In:

Molecular ecology of aquatic microbes (ed. I. Joint), pp. 205–215.

Springer-Verlag, Berlin.

S, R. V., F, R. H. & L, S. D. (1997). Increase in

soluble reactive phosphorus transport in grassland catchments

in response to soil phosphorus accumulation. In Phosphorus loss

from soil to water (ed. H. Tunney, O. T. Carton, P. C. Brookes

and A. E. Johnston), pp. 356–358. CAB International,

Wallingford.

S, M., K, P. & J, E. (1993).

Eight years of internal phosphorus loading and changes in the

sediment profile of Lake Søbygaard, Denmark. Hydrobiologia

253, 345–356.

S, W. C., L, D. P., M, K. W. & S,

M. D. (1977). Use of large submerged chambers to measure

sediment-water exchanges. Water Research 11, 461–464.

S, P. A., C, S. R. & L, R. C. (1997).

Internal phosphorus loading in Lake Mendota: response to

external loads and weather. Canadian Journal of Fisheries and

aquatic Sciences 54, 1883–1893.

S! , I. (1997). A European fertilizer industry view on

phosphorus retention and loss from agricultural soils. In

Phosphorus loss from soil to water (ed. H. Tunney, O. T. Carton,

P. C. Brookes and A. E. Johnston), pp. 311–328. CAB In-

ternational, Wallingford.

S, W. & M, J. J. (1996). Aquatic chemistry: chemical

equilibria and rates in natural waters, 3rd edition. Wiley-Inter-

science, New York.

T, A. W., E, W. M. & S, E. C. (1971).

Nutrients in streams draining woodland and farmland near

Coshocton, Ohio. Water Resources Research 7, 81–90.

T, A. D. F. (1973). The chemistry of phosphorus. (Pergamon

Texts in Inorganic Chemistry, Part 3). Pergamon Press,

Oxford.

T, H., B, A., W, P. J. A. & E,

P. A. I. (1997). Phosphorus fertilizer strategies : present and

future. In Phosphorus loss from soil to water (ed. H. Tunney,

O. T. Carton, P. C. Brookes and A. E. Johnston), pp. 177–

203. CAB International, Wallingford.

T, L., H, H., K, T. & T,

P. (1998). Increased bioavailability of sediment phosphorus

due to silicate enrichment. Water Research 32, 2001–2008.

T, A. J. & P, R. H. (1984). Phosphate exchange

between littoral sediments and overlying water in an

oligotrophic, north temperate lake. Canadian Journal of Fisheries

and aquatic Sciences 41, 1609–1617.

V, R. A. (1968). Scientific fundamentals of the

eutrophication of lakes and flowing waters with particular reference to

nitrogen and phosphorus as factors in eutrophication. OECD, Paris.

V, R. A. (1976). Advances in defining critical

loading levels for phosphorus in lake eutrophication. Memorie

dell’Istituto italiano di Idrobiologia 33, 53–83.

V, R. A. & K, J. (1980). The loading

concept as basis for controlling eutrophication philosophy and

preliminary results of the OECD programme on eutrophi-

cation. Progress in Water Technology 12 (2), 5–38.

W, S., MC, E. & D, J. A. (1992). Sigmoid

relationships between phosphorus, algal biomass and algal

community structure. Canadian Journal of Fisheries and aquatic

Sciences 49, 2605–2610.

W, R. G. (1995). Death, detritus and energy flow in

aquatic ecosystems. Freshwater Biology 35, 83–89.

W, J. D. H., J, J. M. & T, R. L. (1976).

Forms of phosphorus in the surficial sediments of Lake Erie.

Journal of the Fisheries Research Board of Canada 33, 413–429.

W, J. D. H., S, J. K., A, D. E. & H,

R. F. (1971). Characterisation of inorganic phosphate in non-

calcareous lake sediments. Proceedings of the Soil Science Society

of America 35, 556–561.

W, J. D. H., S, J. K. & H, R. F. (1970).

Adsorption and desorption of inorganic phosphorus by lake

sediments in a 0.1M NaCl system. Environmental Science and

Technology 5, 1113–1120.

W, R. J. B. (1976). The chemical composition of rain,


land drainage and borehole water from Rothamsted, Broom’s

Barn, Saxmundham and Woburn Experimental Stations.

Technical Bulletin, Ministry of Agriculture, Fisheries and Food 32,

174–200.

W, J. A. (1977). Contribution of point sources, especially

domestic sewage. In Lake pollution prevention by eutrophication

control (ed. W. K. Downey and G. Ni Uid), pp. 105–110.

National Science Council, Dublin.

W, P. J. A. (1997). Phosphorus loss to water from

agriculture in the UK. In Phosphorus loss from soil to water (ed.

H. Tunney, O. T. Carton, P. C. Brookes and A. E. Johnston),


sources and bioavailability of phosphorus fractions in freshwaters: a british perspective

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