[advances in ecological research] advances in ecological research volume 14 volume 14 || the...

ADVANCES IN ECOLOGICAL RESEARCH, VOLUME 14

The Decomposition of Emergent Macrophytes in Fresh Water

NICHOLAS V. C. POLUNIN

I. Introduction . 11. Sources of Detritus .

A. Primary Resources . B. Secondary Resources . A. Microbial Colonisation . B. Effects of Microorganisms . C. Animal Colonisation .

IV. Decomposition of the Litter . A. Leaching . B. Decomposition . A. Breakdown . B. Microbial Activity . C. Breakdown, Mineralisation, and Particle Production

VI. Chemical Changes in the Litter . A. Major Ionic Constituents . B. Carbon, Nitrogen, and Phosphorus. .

VII. Decomposition of Dissolved and Particulate Matter . A. Suspended and Dissolved Organic Matter . B. Deposits . A. Physicochemical Environment . B. Succession. C. Large Aquatic Consumers . D. Potential Economic Importance .

. References .

111. Colonisation by Aquatic Organisms .

V. Effects of Animals on Litter Decomposition .

VIII. The Swamp Environment .

IX. Summary: Fate of the Detritus

115 116 117 118 119 119 121 122 125 126 127 130 130 133 135 136 136 137 138 139 141 143 143 146 147 148 150 153

I. INTRODUCTION

The emergent macrophytes constitute a significant ecological group among the plants which grow in freshwater habitats. These species include those which are rooted in the sediment, have part of the plant above the water surface for most of the year, and have elongate emergent stems with long cylindrical or narrow flat leaves (Hutchinson, 1975, p. 125). They live characteristically in marginal and temporally variable environments.

115

Copyright 0 1984 by Academic Press, Inc. All rights of reproduction in any form reserved.

ISBN 0-12-013914-6

116 NICHOLAS V. C. POLUNIN

Commonly adapted to function in anaerobic soils, the roots of the plants can make use of a large reservoir of nutrient elements available there (Scul- thorpe, 1967; Bristow, 1975). Production for most of the year is not limited by turbidity of the water because the leaves are in the air above the surface, nor by the availability of water which is nearly always superabundant. These communities are among the most productive known (Brinson et al., 1981). Most of this organic matter can remain ungrazed, however, and the bulk of it typically enters detrital systems (Imhof, 1973).

Unlike those of fully submerged plants, the stems of emergent forms must resist strains from wind and surface water movement, and thus contain much structural fibre (Polisini and Boyd, 1972), which is slow to decompose. The environment of emergent swamps is quite distinctive from that of open water, and in many respects it may not be conducive to decomposition. Purely in terms of carbon flow, emergent macrophytes can be important contributors to the metabolism of fresh waters (Wetzel, 1975; Moss, 1980). The extent to which these plants play a role, however, in functions more significant for the ecosystems of which they are a part than the mere addition of organic matter, such as by regenerating key nutrients or providing a source of trophic energy, is thus far from clear. Probably such roles are modified in intensity and quality by features of the detrital matter itself and of the environment in which it decays.

While knowledge of carbon flux in freshwater swamps has been described in general terms recently (Brinson et al., 1981), I wish to deal here in particular with the processes involved in decomposition. Beginning with a description of the sources of emergent plant organic matter reaching fresh water, I assess what happens to this material, particularly in terms of the agents of decomposition, of how they act, and of how they change, thus influencing the detritus itself, over time. The fate of the dissolved and particulate material is treated separately because its characteristics tend to be different from those of the litter. In Sections V and VIII, among other things, I mention effects of macroinvertebrate animals in some detail, describe distinctive environmental features of emergent macrophyte swamps, and at- tempt to assess how decomposition there may be a significant element in the functioning of freshwater systems.

11. SOURCES OF DETRITUS

In a strict sense, the emergent macrophytes comprise certain members of the Equisetaceae (genus Equisetum), Sparganiaceae (Sparganium), Erio- caulaceae (Eriocaulon), Cyperaceae (Carex, Cladium, Cyperus, Schoene- plectus, Scirpus, Dulichium, Eleocharis), Typhaceae (Typha), Juncaceae (Juncus), and Gramineae (Echinochloa, Glyceria, Hydrochloa, Leersia,

DECOMPOSITION OF MACROPHYTES 117

Paspalum, Zizania, Phragmites), and they are widely distributed in the world (Hutchinson, 1975). Infrequent in streams, they can be common in the lower, slow-flowing portions of rivers (Sirjola, 1969; Westlake, 1973) and abundant in the littoral zone of lakes, at both low and high latitudes (Hutchinson, 1975). In flood plain habitats all over the world emergent macrophytes are conspicuous in seasonally inundated areas and also as permanently floating mats (Welcomme, 1979).

Detritus, the resource of decomposition, can originate from any nonpre- datory loss of organic matter from any trophic level (Wetzel et al., 1972). The primary resources are those derived directly from the plant; they include litter and the soluble matter which may be lost from plants while they are alive. Secondary resources are provided indirectly, for example by the defecation or death of animals feeding on the litter (Swift et al., 1979).

A. Primary Resources The chemical composition of the litter can vary between organs (Buttery

et al., 1965; Gaudet, 1977; Ho, 1979), and within organs according to age and position on the plant (Davis and van der Valk, 1978a). Submerged parts of Phragmites australis (Tobler, 1943) and Scirpus subterminalis (Boon et al., 1982) have a different constitution from aerial parts. Distinct organs may vary considerably in their pattern of fall into fresh water: in P. australis in temperate areas the leaves typically fall in the winter, the stems may stand for a year or more before collapsing, and the rhizomes may start decomposing only after about 3 years (Fiala, 1973).

Closely adjacent plants may vary substantially in their structure (Dy- kyjovA and HradeckA, 1973; Ho, 1980), metabolism (Shaver et al., 1979; Alexander et al., 1980), and chemical constituents (Ho, 1979). In particular, plants from different vegetative clones may have different phenotypes, which may in turn influence the features of their decomposition. The pattern of input may also vary on microgeographic scales. Plants in areas exposed to fire, grazing damage, or physical breakage may senesce earlier in the season than those at undisturbed sites; breakage of the stalk leads to the death of distal parts (Hiirlimann, 1951). Even at a single locality there may be significant variation in the detrital input between years, with changes in senescence pattern and in chemical composition. At the community level, the species composition of emergent assemblages may vary in time and space (Day, 1982), so influencing the general detrital input.

The routes by which the emergent macrophytes contribute to the detrital pool are numerous. Some leaching may occur from live-standing plants, although most attention has been given to the loss of inorganic ions; such losses may be enhanced by aphids feeding on the plant (Szczepadska and Szczepanski, 1973). For most ions, however, such leaching evidently in-


creases at death; many soluble components are released from the dead- standing above-ground parts at this time (Bayly and O’Neill, 1972). Struc- tural materials remain in above-ground parts of the plant, and will enter the water only when the latter fall. The chemical makeup of the litter which is eventually provided to the detrital pool is thus very different from that of the material potentially available after growth has ceased. Leaves and stem fragments commonly float freely on the water to begin with and, particularly at the open-water edge of a reed swamp, may thus not sink at the site where they originally fell. At least in P. australis the latter is notably the case for stems (Roos, 1982).

The main primary resources of decomposition are the various types of litter. The below-ground parts may constitute a large portion of the total production (Bray, 1963; Fiala, 1973; Bernard and Solsky, 1977; den Har- tog, 1978; Klopatek and Stearns, 1978; Alexander et al., 1980), but there is little information on these. Far more than this is known about the input of above-ground parts to detrital systems in fresh water, and this has been particularly easy to study where litter-fall occurs at a single time of the year, as is the case for the leaves of P. australis at higher latitudes.

The pattern of senescence varies considerably between species: in Ca- nadian pothole lakes senescence occurs from early June in Carex atherodes, from late July in Scirpus validus and Scirpusfluviatilis, and from Septem- ber in Sparganium eurycarpum and Typha glauca (Davis and van der Valk, 1978a). In the Carex lacustris stand studied by Bernard and Solsky (1977), shoots remained standing for up to 1 year, but some plants senesced after only 3-4 months. Much mechanical damage can occur, as when geese (KvEt and Hudec, 1971) or cattle (Uotila, 1971) trample the vegetation. Fire may also be important in the dieback of emergent vegetation (Egler, 1952; Ru- descu et al., 1965). At high latitudes at least, premature senescence may be influenced in particular by competition between plants in stands, storms, and insect attack, but these contributions to senescence, and thus to the detrital input, may be small, and most death of above-ground parts begins with the onset of frosts. Other abiotic factors such as ice-scour also influence the production of detritus (Uotila, 197 l), and physical fragmentation occurs from dead-standing plants (Davis and van der Valk, 1978a). Little work has been carried out on emergent macrophytes in the tropics, but evidence suggests that litter can be produced continually (e.g., Gaudet, 1977).

B. Secondary Resources Most attention is given to primary resources in decomposition, but de-

trital matter derived indirectly from emergent macrophytes could often be significant. An array of animals has been observed to feed on emergent


macrophytes (Gajevskaya, 1969), and these plants seem not to possess high levels of compounds which might deter grazers (Hutchinson, 1975). It is, however, widely considered that grazing of the living plant accounts for only a small fraction-usually far less than 20Vo-of the net production (Smirnov, 1961; PelikPn et al., 1971; Imhof, 1973; Dobrowolski, 1973; Skuhravg, 1978; PelikPn et al., 1978). An important factor in this low level of herbivory could be nutritional quality. The limiting concentrations of nitrogen in macrophytes, for example, constitute carbon:nitrogen ratios (Gerloff and Krumbholz, 1966) which are far above those which are probably necessary for the maintenance of grazing animals (Russell-Hunter, 1970). Where the content of nitrogen is enhanced, grazing can greatly increase (cf. Onuf et al., 1977). Aquatic macrophytes also contain high levels of elements such as potassium and phosphorus which may be in short supply in adjacent communities (Howard-Williams and Junk, 1977), and these plants can thus be a valuable source of such substances for animals. Ex- tensive grazing of emergent macrophytes does occur on occasion, such as by insects (Davis and van der Valk, 1978a), fish (Matthes, 1964, reported in Welcomme, 1979; Krzywosz et al., 1980), birds (Dobrowolski, 1973), and mammals (Burgis et al., 1973). Emergent vegetation may also be cropped for ends other than nutrition: animals such as the muskrat use it for con- structing lodges (PClikan et al., 1971). Grazing and other animal damage tend to be highly variable according to plant species, locality, and time, but high levels of these are thought not to be typical of natural emergent swamps, and to be more likely in disturbed sites (Brinson et al., 1981).

111. COLONISATION BY AQUATIC ORGANISMS

The fall of plant litter into the water typicaIly initiates colonisation by aquatic organisms. Patterns of such colonisation are examined here to- gether with effects specifically of microorganisms, while the effects of animals on the litter will be treated later (Section V).

A. Microbial Colonisation Microorganisms such as bacteria are far more abundant on submerged

litter than when it is still attached to the dead-standing plant (Ijlehlovh, 1978a,b). After wetting, bacterial numbers may increase considerably, and commonly reach a peak in a matter of days. OlPh (1972) examined particles of Phragmites detritus and found that microbial biomass as indicated by adenosine triphosphate (ATP) levels reached a maximum at 4-6 days and thereafter declined steadily, to reach a level that was stable for 13 up to 20


days, when observations were completed. In the measurements of Hargrave (1972) on Phragmites particles, oxygen consumption increased for the first 2-3 days, but fell to a fairly stable level by approximately 7 days. Under natural summer conditions, Krashenninikova (1958) noted a peak in bacterial numbers at 25 days on Phragmites litter, and at 35 days on Carex sp. Colonisation of artificially comminuted material could thus be more rapid than intact litter, and the pattern may vary according to plant species. ATP levels on litter of Scirpus acutus also increase with time; at 25°C and under aerobic conditions a peak is reached at about 50 days (Godshalk and Wetzel, 1978a). The ATP level attained on S. acutus litter, however, is greater in aerobic than in anaerobic conditions, and higher at 25 than at 10°C. En- vironmental conditions after submergence thus also influence the pattern of microbial colonisation.

As a plant senesces, there is evidently an increase in the number of fungal species which are able to fructify on experimental agar plates (Apinis et al., 1972); this is probably due to the loss of host plant resistance to such fungi with time. On submerged Phragmites material two stages in microfungal colonisation have been demonstrated (Taligoola et al., 1972). The earlier phase is characterised by the pycnidial form Diplodina, and the persistence of some aerial Hyphomycetes and Ascomycetes. The subsequent stage sees an increase in the microfungus Leptosphaeria culmifraga and a decline in Diplodina; in general the material is now colonised by aquatic Ascomycetes, Sphaeropsidales, and Hyphomycetes among the microfungal forms. Pugh and Mulder (1971) reported changes in the fungal species composition of decaying Typha latifolia leaves, on which nematophagous species become common in later stages. There is therefore evidence of aquatic mycofloral succession on decomposing emergent macrophyte litter, although the mechanisms involved in this context have yet to be studied. It is possible that some microorganisms are responding to change in the detrital resource. Some fungi may be quite plant species specific in the litter which they colonise (e.g., Ingold, 1943, 1951; Ingold and Chapman, 1952), and particular species may be characteristically more abundant on certain plant parts than on others (Pugh and Mulder, 1971). At least in the case of T. latifolia, microfungi are not abundant on the rhizomes or roots of the plant (Pugh and Mulder, 1971), although evidence of fungal succession on pieces of the root of Iris pseudacorus has been presented (M. W. Assawah, 1956, reported in Chesters, 1960). Direct interactions between microbial decomposers may also contribute to successional patterns.

Even less has been noted in the literature about the nature of the bacterial flora on emergent macrophyte litter in fresh water, although these microorganisms may often be in evidence more than the fungi and Actinomycetes (Gorbunov, 1953; Pieczyhska, 1975). The work of Olhh (1972) on Phrag-


mites particles has suggested that a succession of bacterial forms may colonise the detritus. The bacteria eventually decline considerably as shown by the fact that their numbers and biomass per unit weight of detritus are far lower on highly degraded particles from reed beds than on freshly submerged litter (Ulehlovii, 1978a,b), and by the fact that oxygen uptake also declines with time (Hargrave, 1972). Comparison of “young” with experimentally matured Phragmites litter has suggested that, early on, microbial activity could be limited by the availability of elements such as nitrogen and phosphorus, whereas later in decomposition the availability of labile substrates is more likely to be influential (Polunin, 1982b).

This respiration is not all attributable to decomposer species (Mason and Bryant, 1975a; Triska and Sedell, 1976); significant microbial activity may be detected on inert surfaces (Hargrave, 1972; Bobbie et al., 1978). Indexes of microbial activity such as oxygen consumption may thus give a poor measure of microbial decomposition of the immediate detrital resource; particularly early in decay, much of the increase in bacterial numbers may depend on the release of soluble substances by abiotic processes such as leaching (Section IV,A). Where light is sufficiently intense, diatoms and other microalgae may also be significant microbial colonisers of the litter. If silica is limiting, the diatoms may benefit from the supply of this element afforded by grasses such as Phragmites (Jmgensen, 1957). The possibility exists that such nondecomposers interfere with microbial decomposers by competing for nutrients or space or in other respects (Vaatanen and Sund- quist, 1977), but such interaction has not apparently been explored, although the experiments of Mason (1976) have indicated some antagonism between fungi and bacteria susceptible to antibiotics. Periphyton algae, protozoa, nematodes, and other microbiota may account for one-third or more of the oxyen uptake of Phragmites litter early in decay (Mason, 1976).

B. Effects of Microorganisms Research on reed beds in the Austrian Neusiedlersee has indicated that

microorganisms, in trophic energetic terms, may be of overriding importance in emergent litter mineralisation (Imhof, 1973). Although this seems likely to be the case more widely, the necessary analytical work has not been reported from elsewhere. Mechanisms and rates of decomposition are examined below (Section IV).

Bacteria and fungi may be equally important specifically in cellulolysis (Zdanowski, 1977), but the experiments of Mason (1976) with Phragmites leaf litter as a whole suggest that bacteria may become more important than fungi within 4 months’ decomposition, although the two groups of decomposers may be playing an equal role by the end of the first month. Obser-


vations on tree leaves in freshwater streams have suggested a comparable shift in the dominant microbial decomposers with time, and raised the possibility that the fungi may somehow make the substrate more susceptible to bacterial colonisation (Suberkropp et al., 1976a). In a similar fashion decomposer microorganisms make the substrate more acceptable to macroinvertebrate detritivores as food (Gorbunov, 1953), an effect that can evidently be mimicked by treatment with hot acid (Barlocher and Kendrick, 1975a). The microorganisms themselves, however, are of far greater nutritional value to detritivores than the leaf litter alone, for the animals digest the latter inefficiently (see below Section V,A).

Apart from decomposing the substrate, such litter microorganisms may have many indirect effects on the systems of which they are a part. Partic- ularly in areas with a poor oxygen supply, they may for example lead to substantial depletion of oxygen (e.g., Gorbunov, 1953), and this in turn affects the distribution and abundance of other organisms. Swamps commonly present harsh conditions for organisms which respire aerobically. Another respect in which the decomposer microorganisms are particularly influential is in the recycling of key nutrients. The high turnover of available forms of nitrogen and phosphorus to which they contribute may be important to the macrophytes themselves, and also to organisms living at the fringe of the swamp. In addition to making the detritus more amenable to detritivore feeding, the microbial decomposers also convert detrital organic matter into their own biomass, a form in which the material is more accessible to macroinvertebrates. These microorganisms tend to be far more abundant around emergent macrophyte beds, where the input of organic matter is intense, than in open water (Krashenninikova, 1958), and may constitute an important source of food in the littoral (see Section VIII).

C. Animal Colonisation Wide variation in the aquatic species which colonise decomposing emer-

gent macrophyte litter occurs because the assemblages can differ greatly over small distances (Section V111,C). The relative abundance of macroinvertebrates colonising bags with emergent litter in them can vary considerably between lakes (Dane11 and Andersson, 1982), from one site to another within a lake, and over short vertical distances (N. Polunin, unpublished).

At a single locality there may also be large quantitative and qualitative changes in the macroinvertebrates colonising emergent litter over time. For Phragmites australis leaf litter placed in 10-mm-mesh bags in the main brick pond at Wicken Fen, Cambridgeshire, at the end of January, a maximum in numbers of individuals was indicated in the first half of the summer; at the observed peak, the gammarid Crangonyx pseudogracilis was dominant


in terms of numbers and weight, and the isopod Asellus meridianus was the second most important single species (Fig. 1). Starting in December, Webster and Simmons (1978) found a buildup of macroinvertebrate numbers on T. latifolia litter up to the ensuing September, By that time the litter had been skeletonised, and macroinvertebrate numbers became greater on litter of other plant species. Macroinvertebrate taxa which at any stage accounted for more than 5 % of all individuals included the gammarid Hy- alella azteca, the isopod Asellus sp., chironomids, flatworms, and oligochaetes. H. azteca was dominant in numbers in the spring; Asellus

Am= A g t y ~ ~ ~ p q etana AMD = Asellus meridianus BTH = Bithynia spp. CAE = spp. CHI = Midge larvae

HIR Leeches PLN = Planorbis $p. I 1 = Others

CRX = Crangonyx pseudgnv ilis

L

0 t -

MONTH

Fig. 1. Changes in the total abundance and composition of the macroinvertebrate fauna colonising Phragrnites australis leaf litter enclosed in 10-mm-mesh bags at the edge of the reed swamp in the main brick pond of Wicken Fen, Cambridgeshire, England. The experiment began on 31 January 1977, and the densities, expressed as numbers of individuals per gram of dry litter, were corrected for animal colonisation of blank bags which contained no litter. From N. Polunin (unpublished).


dominated in early summer, and was exceeded in numbers by the chironomids toward the end of the summer. Starting in May, Danell and Sjoberg (1979) observed a peak in macroinvertebrate populations on litter of Carex spp. and of Equisetumfluviatile at approximately 5 months, although there was not necessarily a steady increase up to that point. At the observed peak the isopod Asellus was dominant in terms of biomass. There is thus a broad similarity in the assemblage of litter-colonising macroinvertebrates at a range of localities. There may, however, be local differences according to community and species of litter. Macroinvertebrate biomass is greater on Carex than on Equisetum (Danell and Sjoberg, 1979) or Typha (Danell and An- dersson, 1982) litter; the latter authors also record qualitative and quantitative differences on the same litter in separate lakes.

While the number of macroinvertebrate individuals increases with time, so also does the number of species (Fig. l), but the role which any change in the litter material itself plays in such changes has been little examined. When colonisation of Phragmites leaf litter that is either fresh or has been allowed to decompose naturally in situ is compared, only some of the macroinvertebrate taxa which increase with time exhibit significantly greater numbers on the mature litter (N. Polunin, unpublished). Much of the macroinvertebrate increase in abundance at such localities may just be a seasonal growth in populations (Fig. 2). In their experiments on macroinvertebrate colonisation, Webster and Simmons (1 978) compared containers with either Typha litter or plastic strips, and found that although there were more animals on the real litter at two sampling times in spring, the three subsequent samples yielded no difference between the two substrates. Street and Titmus (1982) have found that variations in the structural properties of the material placed in litter bags can have 14 times the influence that the presence or absence of added detritus has on total macroinvertebrate numbers. Similar evidence for the “inert” role of plants and plant litter has come from other studies in lentic (Macan and Kitching, 1972; Soszka, 1975; Voshell and Simmons, 1977) and lotic (Egglishaw, 1964; Winterbourn, 1978) environments. Much of the macroinvertebrate colonisation observed may thus be related to factors other than changes in the litter itself over time.

For the longer term, Opalinski (1971) has examined the macroinvertebrates colonising the submerged parts of new and old stems of Phragmites australis. The young stems tended to support more individuals and more species. Several taxa were absent from the old stems and these included oligochaetes and the mollusc Dreisena polymorpha. Leeches, caddis larvae, and flatworms were more abundant on the new stems than on the old (Opal- inski, 1971).


a

Fig. 2. (a) Diagram of a transect across a Curex- and Glyceriu-dominated stand on the southern, forested shore of Radov Pond, southwestern Bohemia, Czechoslovakia, showing the position of three stations (A, B, C) at which hand-grab samples were made. The water levels at 5 April and 22 September are shown. (b) Macroinvertebrate biomass estimates were taken in spring (ApriVMay), early summer (June/July), and late summer (AugustISeptember). Solid bars, carnivores; open bars, noncarnivores. After Dvoihk (1970b).

IV. DECOMPOSITION OF THE LITTER

Weight loss from the litter typically exhibits a high initial rate, which subsequently declines. Although it is difficult experimentally to separate the two processes, the initial rapid decomposition is primarily a case of leaching, the abiotic loss of soluble components. The later phase depends pre- dominantly on the activities of decomposer organisms and is generally assessed as a breakdown rate, the speed at which weight is lost from contained or tethered litter regardless of its fate.


A. Leaching

The loss of soluble matter through leaching is particularly significant in aquatic habitats because the material is in continual contact with water (Brinson et al., 1981). While decay in air may be negligible (Gorbunov, 1953; Hobbie et al., 1980), on entering the water, the litter swiftly loses weight. Using mercuric chloride as an inhibitor of microbial activity, over 50% of the weight loss from Carex aquatilis can be shown to be due to “leaching” (Hobbie et al., 1980), and using sodium a i d e on Phragmites leaf litter, 80% of the weight loss up to 20 days is attributable to “leaching” (Polunin, 1982b). The latter figure is comparable to that of Harrison and Mann (1975) for leaves of the sea grass Zostera.

The rate of early weight loss may be enhanced by reducing the size of particles (Szczepanski, 1978; Larsen, 1982; Polunin, 1982b), and also by raising the temperature (Paul et al., 1978). The weight loss of Phragmites leaf litter after 8 days at 26°C is twice (17.5%) that at 4°C (8.8%; N. Po- lunin, unpublished). The leaching of ions appears little changed by oxygen depletion, and increasing carbon dioxide affects only the loss of calcium, which is accelerated (Planter, 1970). Comparing ionic elution into natural and distilled water, rates for phosphate, ammonia, sodium, and potassium remain similar, but levels of calcium and total conductivity of the medium are lower in lake water than in deionised water (Planter, 1970). The extent of mineral ion loss from dead-standing culms of Cyperus papyrus may be influenced by the amount of rainfall and also by its pH (Gaudet, 1977). A possible effect of pH on the release of humic substances has been suggested by Howard-Williams and Howard-Williams (1978). In conclusion, although microbial activity may increase greatly in the first few days of submergence, when biotic inhibitors are added or the environment is experimentally al- tered, a large early weight change still occurs. The exact mechanism is unclear, but microorganisms apparently make little contribution to this loss, which may be substantial.

Much of the initial leaching from emergent macrophyte litter occurs in the first few days. With dead shoots of Typha domingensis most of the increase in water conductivity, following the loss of ions, happens in the first day (Howard-Williams and Howard-Williams, 1978). This rate is similar to that found by Planter (1970) for Phragmites australis litter, although OlAh (1972), working with fine particles of P. australis, suggested a longer leaching time. Planter (1970) observed that the greatest increase in water conductivity with P. australis litter occurred in the first 50 minutes. Dane11 and Sjoberg (1979) reported a 6.8% weight loss from Equisetum litter within 1 hour, although Carex litter lost only 0.9% in the same time.

To begin with, soluble ionic constituents make a substantial contribution to the leachate, although different ions are eluted at characteristic rates.


Important features of this process are examined in Section VI. Whereas much attention has been paid to the release of minerals, little work has been carried out on the release of organic matter. Otsuki and Wetzel (1974) observed that most of the soluble organic matter released from the submerged macrophyte Scirpus subterminalis was lost from the litter in the first 5 days. Howard-Williams and Howard-Williams (1 978) observed that the release of humic compounds from T. domingensis was rapid in the first 5 days, and thereafter slower. While it has not been measured specifically, the release of soluble organic matter at this stage can evidently be substantial, although this may depend on the plant part involved: the total leaching loss from leaves is some five times greater than that of stems in P. australis (Polunin, 1982b). Larsen and Schierup (1981) did not detect an obvious leaching phase for the Phragmites litter which they studied, and they suggested that such leaching reported in other cases was an artifact of oven-drying at high temperature. Initial rates of weight loss can be modified by oven-drying, but in Cambridgeshire fenland P. australis leaching is not an artifact. Leaching may thus be important in the release of soluble ions and organic compounds from the litter.

B. Decomposition Decomposition has only been crudely assessed as a weight loss from

coarse-mesh bags in most cases, and usually includes the leaching phase. Some results on such breakdown rates for emergent macrophytes have recently been summarised by Brinson et al. (1981), and are presented as a decay constant k, which is estimated as the slope of the regression of In X,/X against time, according to a first-order decay model. Reported values of k cover a large range from 0.34 year - for Scirpusfluviatilis to 2.03 year - for Scirpus validusin central North America. Some higher values have been reported, for example for Typha IatIfoIia (Webster and Simmons, 1978; Hill and Webster, 1982), but in conclusion the litter commonly takes at least 1 year to disappear from litter bags. An important point, however, is that there are large differences in the methods used, and often precise data on the litter material and prevailing natural conditions are not available. Although a systematic explanation of this range of rates is therefore far from possible, a number of factors are known to influence decomposition rates, and some of these are reviewed here.

From experiments in which decay has been measured in a standard manner at the same locality (e.g., Odum and Heywood, 1978; Davis and van der Valk, 1978a; Dane11 and Sjoberg, 1979; Blake, 1982; Hill and Webster, 1982), it is clear that the pattern and rate of weight loss can vary significantly between species. There are thus factors intrinsic to the litter, and characteristic of species, which are influencing decomposition. The extent


of leaching is determined largely by the content of matter which is readily lost into solution. The biotic processes usually make a larger contribution to decomposition than does leaching, and factors influencing these thus have a larger effect on decomposition as a whole. Of the intrinsic factors, two types are at present considered to play an overriding role in limiting decomposition. These are structural compounds that are resistant to decomposer action, and second, key elements that determine the value of the detritus as food.

Gorbunov (1953) noticed that macrophytes with high lignin levels, such as Phragmites and Sparganium, were slowest to decay. In P. australis leaf litter, lignin tends to increase in concentration over time (Polunin, 1982b), and Sircar et al. (1940) observed that rice-straw lignin was one of the slowest components of the litter to decompose. After comparing decomposition rates of a number of freshwater macrophytes, Godshalk and Wetzel(1978a) concluded that the total fibre content of particulate detritus did adversely affect its decay. Under a range of four experimental conditions of varying aeration and temperature, however, structural compounds including lignin, cellulose, and hemicellulose often remained constant or decreased in concentration in the detritus. There was thus little evidence in this case that more labile components than these were selectively removed by microorganisms (Godshalk and Wetzel, 1978a). Crawford and Crawford (1976) have shown that lignin derived from Typha can be mineralised quite rapidly in the soil. The decomposition of Carex-derived lignin, however, is not accelerated by adding nutrients, although that of cellulose is (Federle and Ves- tal, 1980b).

Aquatic microorganisms on emergent macrophyte litter tend to absorb available forms of phosphate and nitrogen (Gorbunov, 1953; Godshalk and Wetzel, 1978a; and see also Section V1,B); the supply of these elements thus influences decomposition. Addition of nitrate and phosphate to P. australis litter increases microbial aerobic respiration and actual decomposition (Po- lunin, 1982b). At artificially nutrient-rich sites, P. australis litter may take up oxygen and decompose more rapidly than at nutrient-poor sites (God- lewska-Lipowa, 1975), although in the two lakes compared by Andersen (1978) the increase in microbial oxygen consumption was far greater at the eutrophic site (approximately twice) than was the increase in rate of weight loss (approximately 10%).

Also, within species there may be changes in the decayability of the litter collected in different years at the same site (N. Polunin, unpublished), and different plant parts decompose at varying rates. P. australis leaves take up oxygen (Andersen, 1978) and decompose fllehlovd, 1978b; Larsen and Schierup, 1981; Blake, 1982; Polunin, 1982b) at two to seven times the rate of P. australis stems per unit weight.


When the same litter is placed at different localities, varying decomposition rates are commonly found; there are thus also environmental influences on decomposition, as well as factors intrinsic to the litter. Changing rates of decomposition may for example be found along vertical gradients from land toward deep water. Typically, permanently submerged macrophyte litter is broken down far more rapidly than litter which is intermit- tently inundated or kept largely dry (Gorbunov, 1953; Boyd, 1970; Pieczyriska, 1972; Brinson, 1977; Odum and Heywood, 1978; Furtado and Verghese, 1981; Hill and Webster, 1982). Seasonal cycles of wetting and drying are characteristic of most emergent macrophyte communities, but as yet there appears to be no simple relation between the extent of decomposition and frequency or length of flooding periods (Brinson et al., 1981). Breakdown rates may be greater in open water outside reed beds than inside them (Gorbunov, 1953); reasons for this may be associated with lower light levels (Blake, 1982) and higher acidity (McKinley and Vestal, 1982) inside the reed swamp. Below the littoral, decomposition rates commonly decrease with depth (Federle and Vestal, 1980a). This may be related to vertical changes in environment, whereby oxygen levels and temperature decrease at depth (Godshalk and Wetzel, 1977; Reed, 1979; Polunin, 1982a). Such environmental changes influence the activity of microorganisms and large detritivores and, possibly, abiotic processes such as leaching. There has, however, been no detailed study precisely to explain such depth patterns.

Apart from these vertical influences on decomposition there are also factors which vary horizontally. Litter of both Carex and Equisetum loses weight more rapidly in Carex beds than it does in Equisetum beds (Dane11 and Sjoberg, 1979), although the reasons for this have not been ascertained. Carex litter disappeared more rapidly in a neutral pH, nutrient-rich Swedish lake than in an acid, nutrient-poor one (Dane11 and Anderson, 1982), but again the cause of this relationship has not been elucidated. There may thus be habitat-related influences on decomposition of the same material, although variations in the plant species composition may often be more influential in determining local differences in decay patterns (Day, 1982). Seasonal effects at single sites may also be important: Brinson (1977) showed that monthly decomposition rates of Nyssa aquatica leaves exhibited a high positive correlation with maximum and minimum monthly soil temperatures. For P. australis Ulehlovh (1978b) found that litter placed in ponds in the winter decayed much more slowly than summer litter samples, while Polunin (1 982b) showed that decomposition rates, excluding initial leaching, and starting in the winter, tend to be instantaneously higher in the ensuing summer than they were initially for the same litter in the winter.

Effects of important environmental factors have been studied more sys- tematically in experimental systems. Over the span of 180 days in aerobic


microcosms, Scirpus acutus detritus decomposes more than twice as fast at 25°C in comparison to 1O"C, and in comparison to an anaerobic environment at both temperatures (Godshalk and Wetzel, 1978b). Excluding the initial 8 days' weight loss, over the subsequent 60 days P. australis leaf litter decomposes more than twice as fast at 16 as at 4"C, but the rate at 26°C is scarcely greater than at 16°C; oxygen consumption of this litter, however, more than trebles at 16°C and approximately quadruples at 26°C relative to 4°C (Polunin, 1982b). With tree leaves, Ward and Cummins (1979) found that there was no significant increase of ATP-estimated microbial biomass with rising temperature, but that microbial respiration did increase with it. The effect of inorganic nutrients such as nitrate and phosphate on decomposition has been noted above. Davis and van der Valk (1978a) found evidence for a positive correlation between decomposition rates and initial nitrogen concentration for a number of emergent macrophytes in north Iowa lakes. Other such comparisons have been attempted, but often over a range of types of macrophyte whose decomposition characteristics are quite distinct. All the same, it is clear that the nitrogen requirement for decomposition may be contributed to by both exogenous and intrinsic sources of the element.

V. EFFECTS OF ANIMALS ON LITTER DECOMPOSITION

Animals such as benthic macroinvertebrates often ingest large quantities of plant detritus (Monakov, 1972; Ladle, 1974; Hodkinson, 1975b). The distribution of detritus in freshwater habitats can influence assemblages of these animals (Egglishaw, 1964). Do animals actually influence decomposition, and if so how? How does such detritivory contribute to the metabolism of the consumer?

Broad studies of some freshwater ecosystems have indicated a negligible energetic role for the larger animals there (cf. Saunders, 1976). It is nevertheless possible that where they are abundant, animals could be contributing to decomposition in ways that would not be recognised in conventional energetic studies. Such effects could be exerted through breakdown, direct and indirect modification of mineralisation, or all three.

A. Breakdown Experimental field manipulations used to investigate animal effects on

breakdown are usually uncontrolled. In particular, different mesh-sized bags employed to exclude certain animals in the field often fail to examine non-


animal factors which may influence results. Even where a significant effect of mesh size is found, it is not always clear that this is attributable to the effect of animals. In one case, a significant mesh size effect observed in shallow water, where macroinvertebrates were abundant, was paralleled by a comparable effect in a deep-water oxygen-depleted site where litter- feeding animals were absent (N. Polunin, unpublished). The variance of breakdown rates is often large, so that any low-level but significant effect of animals may be hard to detect (Mason and Bryant, 1975a). A significant effect of mesh size may also occur because microbial decomposition is reduced in fine-mesh bags, where conditions of aeration and nutrient availability may be less favourable than in coarse-mesh bags.

One way around these deficiencies is to assess effects of animal activities independently. Observations on natural litter in situ, for example, can provide information on animal breakdown if particular types of animal feeding can be identified (N. Polunin, unpublished). In this case an estimate of visible animal grazing was possible only as long as enough of the original leaf remained to assess the starting weight; the problem of assessing levels of detritivore activity after long time periods remains.

Laboratory experiments have enabled the effects of single species or groups of species to be examined far more precisely, but there is still un- certainty in extrapolating these correctly to the field. The combination of studies in situ and in the laboratory has shown that when the right kinds of animals and conditions are provided, breakdown can be enhanced. Such animal feeding often increases with time, either in a somewhat gradual manner (Herbst, 1982), or in a more rapid way after a period of little activity (Anderson and Grafius, 1975; Barlocher and Kendrick, 1975a). Grazing damage on natural Phragmites leaf litter can increase markedly with time, but the reasons for this have been little examined. Colonisation by microorganisms may enhance the nutritive value of the material. Decomposing emergent plant material is more readily fed upon by macroinvertebrates such as the snail Lyrnnaea than material from the live plant (Kolodziejczyk and Martynuska, 1980). Under controlled conditions such feeding increases markedly with temperature, as shown for tree leaves by Anderson and Graf- ius (1975) and Winterbourn and Davis (1976), and for Phragmites leaves by Polunin (1982b). The effect of macroinvertebrates on breakdown may vary significantly with type of litter (Herbst, 1982) and species of macroinvertebrate (Polunin, 1982b).

Although commonly feeding on plant litter, freshwater macroinvertebrate detritivores generally assimilate only one-tenth to one-third of the material which they ingest. Only 14-18% and 22% of leaf protein potentially available were assimilated by Gammams pseudolimnaeus (Biirlocher and Kendrick, 1975b) and Hyalella azteca (Hargrave, 1970b), respectively.


Mason and Bryant (1975a) report an assimilation efficiency of 17% for Gammams pulex fed on Phragmites leaves, while 23% of tree leaves ingested by Asellus aquaticus are assimilated by it (Adcock, 1982). In the isopod A . aquaticus, certain components of the litter may be more efficiently assimilated than others: Prus (1976) reports efficiencies of 9% for lignin, 32.4% for crude protein, 54.1% for lipids, and 72.3% for carbohydrates of alder leaves. The gammarid H. azteca (Hargrave, 1970b) evidently does not assimilate labelled lignin and cellulose at all. Structural compounds such as these can constitute a high proportion of the litter, and yet a lot of macroinvertebrates do not have the ability to digest them (Bjar- nov, 1972). Even where cellulases are present, as in many freshwater molluscs and crustaceans (Calow and Calow, 1975; Monk, 1976), the effect of such enzymes. on the cell wall material in which their specific substrates occur may be negligible (Monk, 1977).

In contrast to their poor capacity to process major components of the litter proper, macroinvertebrates can digest and assimilate microorganisms on the detritus much more effectively. Barlocher and Kendrick (1975b) reported that G. pseudolimnaeus assimilates 90% of the fungal protein which it ingests (in contrast to only 14-18% of the leaf proper), while H. azteca assimilates most microflora with an efficiency of over 50% (Hargrave, 1970b). Detritivores often prefer litter which has been colonised by microorganisms (Calow, 1974) to uncolonised litter, and there is evidence that particular types of microorganisms may be preferred (Gorbunov, 1953; Barlocher and Kendrick, 1976). Hargrave (1970b) has suggested that these animals may selectively ingest those types of food which can be most efficiently assimilated.

Of the organic matter assimilated, only a rather small proportion is probably used in growth (Adcock, 1982). The efficiency with which this is so can vary with species of plant food (Smirnov, 1962), but certain chironomids can survive and grow with cellulose as the sole trophic source, provided that it is colonised by bacteria (Gorbunov, 1953); on sterile detritus these animals die. Growth may vary with substrate quality, for example in relation to nitrogen content (Iversen, 1974), or to microbial levels and activity (Ward and Cummins, 1979) on the litter. Certain freshwater macroinvertebrates are thus able to survive, and sometimes to grow and reproduce on natural litter, as shown for chironomids on tree leaves (Ward and Cummins, 1979), and as suggested for the snails Bithynia tentaculata and Planorbis carinatus, the isopod Asellus meridianus, and larvae of the caddis Limnephilus marmoratus on Phragmites australis leaf litter (Po- lunin, 1982b). Two caddis larvae have been shown to be able to grow on leaves of the tree Alnus, but not on those of Pseudotsuga (Anderson and Grafius, 1975).


It seems likely that the level of detritivore breakdown may often be greatly influenced by predation. Experiments in which large animals have been ex- cluded or enclosed have suggested that predation substantially reduces the biomass of many small species. Fish, for example, may utilise a large proportion of the available secondary production in two northern Polish lakes (Kajak, 1972), although the effects on detrital decomposition are unknown. The role of animals in breakdown over whole swamps has yet to be estimated also. Most work to date has been confined to areas where detritivores occur naturally, and yet it is evident, for example, that there are large areas in swamps from which such animals are often largely absent.

B. Microbial Activity It has been observed that the number of decomposer microorganisms is

reduced by grazing animals (Barsdate et al., 1974; Bryant, et al., 1982). Grazers may also qualitatively alter the microbial community colonising surfaces such as leaves: fungal hyphae may become less conspicuous and bacteria may be more strongly attached than when not grazed (Morrison and White, 1980).

Although microbial numbers can be reduced by grazing, microbial activity can be stimulated by it, and this is not unexpected in situations in which populations of microorganisms are limited by factors such as the availability of primary nutrients. Barsdate et al. (1974) found that bacteria on Carex particles grazed by the protozoan Tetrahymena recycled phosphorus at a greater rate when grazed than when not grazed; the recycling of phosphate was far faster than, and was not dependent on, mineralisation of the substrate. The oxygen uptake by microorganisms on Phragmites litter can be enhanced by snails such as Lymnaea pereger (Mason and Bryant, 1975a) and P. carinatus (Polunin, 1982c), but this stimulation may not happen with the amphipod G. pulex or the isopod A. aquaticus (Mason and Bryant, 1975a). Such effects may also vary considerably with a single species under different conditions: in laboratory experiments, both P. carinatus and A . meridianus tend to reduce litter microbial oxygen uptake and decomposition relative to ungrazed controls at 4"C, but microbial respiration can be enhanced by these detritivores at higher temperatures (Polunin, 1982b).

Some of this increased activity may be due to nondecomposer microorganisms, and enhancement through the release of labile dissolved organic matter by the grazer could occur (Hargrave, 1970a). Some cases of grazer- stimulated microbial decomposition, however, have been demonstrated (Fenchel and Harrison, 1976; Morrison and White, 1980; Fish and Car- penter, 1982; Polunin, 1982c), and this can be substrate specific (Sherr et al., 1982). The mechanisms by which such enhancement of decomposi-


tion occurs through detritivore activities could be many. In microbial communities, inorganic nutrient release from protozoan grazers may be small relative to that from the bacteria themselves; grazing leads to increased uptake and release rates of primary nutrients (Barsdate et uf., 1974). The enhanced recycling can also occur with larger detritivores such as snails, but here the grazers themselves could be releasing significant amounts of such nutrients (Polunin, 1982~). Further, mechanical rasping of the detritus may expose new surfaces to microbial colonisation (Hargrave, 1970a; Mason and Bryant, 1975a), and grazers may beneficially influence the immediate environment, for example by increasing microturbulence (Hargrave, 1970a; Barsdate et uf., 1974). The activity of nondecomposers such as diatoms may also be enhanced on litter surfaces; evidence for this has been found using experimental enclosures in a Cambridgeshire fenland pond (N. Polunin, unpublished).

The low assimilation efficiency of detritivores fed upon natural detritus means that most of the material ingested by detritivores is returned to the system as faeces. Since detrital components susceptible to digestion are selectively removed, the concentration of structural compounds resistant to decomposition tends to be greater than in the original material consumed (Anderson and Grafius, 1975). Nevertheless, the indication is that decomposition of faeces produced by macroinvertebrates may be faster initially than that of the litter fed upon. Kdodziejczyk and Martynuska (1980) found that the snail Lymnueu stagnafis feeding on litter from emergent plants such as Gfyceriu uquaticu produced faeces which lost 50% of their weight in 2 days while ungrazed plant litter required 7-14 days to decay by the same amount. These faeces, however, did not lose any further weight between days 2 and 5 , when measurements were completed. Similarly, both Har- grave (1972) and Mason and Bryant (1975a) report that the faeces of the snail L. pereger can consume oxygen more rapidly than the Phrugmites litter on which they are fed, but although microbial activity up to 3 days is enhanced, later on, by about 5 days, oxygen uptake may return to the rate of the ungrazed litter (Hargrave, 1972, 1976).

The means by which enhanced faecal decomposition relative to the ingested controls may occur is unclear. Mechanically comminuting litter does not necessarily speed decay (Kaushik and Hynes, 1971; Polunin, 1982b), although among fine-grained material the decomposition rate can be inversely related to particle size (Hargrave, 1972; Harrison and Mann, 1975; Polunin, 1982b). Much natural detrital material is but sparsely colonised by decomposer microorganisms, however, and the surface area per unit weight of the material does not appear per se to be limiting (Hargrave, 1972, 1976; Wiebe and Pomeroy, 1972).

It is evident from all such work that while the ultimate source of organic


matter in the system is for the most part detrital, this material is nevertheless colonised by microorganisms which can also be influenced by grazers. This interaction with microorganisms means that detrital systems are potentially far more complicated than they would otherwise appear to be.

C. Breakdown, Mineralisation, and Particle Production

Where breakdown is enhanced by detritivores in situ, the effect on actual mineralisation remains far from clear. If detritivores remove material from the litter, some of it may be lost as “pseudofaeces” (particles broken from the litter but not ingested), or it may be ingested, and most of it, commonly far more than two-thirds, is voided as faeces. Microbial decomposition of a portion of this egested material may be increased (Hargrave, 1976), but whether this portion is a particular type of compound, or a whole range of them which are simply made more accessible by mechanical action, is un- certain. The fate of the breakdown products remains unknown. In short- term laboratory experiments with P. australis leaf litter under favourable conditions, most of the material broken down decomposes, and only 10- 25% of it accumulates in the system. Detritivores which are particularly effective at shredding the litter, however, can produce a higher proportion of particles which tend to accumulate, and thus reduce the amount of material broken down which decomposes (Polunin, 1982b). Such observations emphasise that while breakdown can be enhanced by animals, most mineralisation is dependent on microbial processes.

Animal feeding may increase substantially in response to rising temperature, so that particle production by detritivores may be especially marked in the summer at higher latitudes. Of the finer particulate breakdown products, some may be suspended, and a larger portion of detrital decomposition may thus occur in the water column in the presence of detritivores than without them (Mason and Bryant, 1975a). Some of this suspended material may be used by filter feeders, while in the presence of sufficient water movement some export of organic matter is likely to occur (Polunin, 1982a; Wal- lace et al., 1982). The coarser material produced by animal feeding tends to contribute to the sediment in situ. Production of faecal particulate matter by the snail L. stagnalis alone in reed beds may be high (Kdodziejczyk and Martynuska, 1980), and such material may also be utilised by other animals (e.g., Brown, 1961; Minckley, 1963, reported in Barlocher and Kendrick, 1975b). In trophic-energetic terms the role of macroinvertebrates is a small one (Imhof, 1973), but detritivores could be making important local contributions to the system in other respects. The fate of dissolved, suspended, and deposited organic matter is examined in greater detail in Section VII.


VI. CHEMICAL CHANGES IN THE LITTER

A. Major Ionic Constituents

Many mobile ionic constituents are readily lost from the above-ground organs of the plant on senescence. These ions either may be largely trans- located to below-ground parts, to be stored in the rhizome, or may be mostly lost, temporarily at least, into the water. In the Typha glauca plants studied by Bayly and O’Neill (1972), calcium was lost from above-ground organs at the end of the season, but there was no increase of this element in the rhizomes, and much of it was presumably released from the plant. A similar pattern was inferred for magnesium, and also sodium and iron, although for the latter two no evidence of increased level in the soil was found. Con- versely, potassium and phosphorus were probably conserved by the plant.

Rapid early changes in ionic constituents also occur in the fallen litter. In Phragmites australis leaves, rapid loss of sodium, potassium, magnesium, and phosphorus takes place within the first month, but calcium may increase or decrease in P. australis and Typha latifolia (Mason and Bryant, 1975a; Ulehlovh, 1978b; Polunin, 1982b). For the latter species, Boyd (1970) reported that nearly all magnesium, potassium, and sodium was lost in the first 20 days of decomposition. In fact, for P. australis litter Planter (1970) concluded that most sodium was leached in a matter of minutes, potassium and phosphate in a few hours, and ammonium in a few days, while calcium still increased in the medium after 30 days. With dead-standing Cyperus papyrus culms, sodium and potassium are eluted most quickly, while losses of iron, magnesium, nitrogen, phosphorus, and calcium are slower (Gau- det, 1977). Potassium and phosphate are lost swiftly from littoral grasses of Lake Kariba, while nitrate, calcium, and magnesium are lost more grad- ually (McLachlan, 197 1). Howard-Williams and Howard-Williams (1978) estimated that 26% of the standing stock of sodium, some 12% of the magnesium, 9% of the calcium, and 8% of the potassium were leached from Typha domingensis litter with flooding in Lake Chilwa. In their detailed study of various emergent species, Davis and van der Valk (1978a) reported three categories of nutrient release pattern, concluding that potassium, sodium, and magnesium were lost most readily, that aluminium and iron were lost most slowly, and that calcium tended to be intermediate in its behaviour between these extremes. There was in fact a net accumulation of iron and aluminium in both the Typha and Scirpus after 525 days, while magnesium and calcium also accumulated in the Scirpus (Davis and van der Valk, 1978b). Heavy metals such as zinc, lead, and cadmium tended to increase in concentration in the P. australis litter of two Danish lakes, and lead was apparently accumulated by the litter (Larsen and Schierup, 1981).


B. Carbon, Nitrogen, and Phosphorus An early rapid loss of soluble forms of nitrogen is characteristic of most

emergent macrophyte litters which have been studied. An initial fall in the nitrogen concentration in the litter has been noted, for example, in P. australis (Mason and Bryant, 1975a; Polunin, 1982b), in Carex spp. (Chamie and Richardson, 1978), in T. domingensis (Howard-Williams and Howard- Williams, 1978), and in dead-standing Cyperus papyrus (Gaudet, 1977). In the long-term, however, nitrogen levels may increase, as shown for P. australis (Mason and Bryant, 1975a; Ulehlovi, 1978b; Polunin, 1982b), Carex spp. (Chamie and Richardson, 1978), Deschampsia caespitosa (Hodkinson, 1975a), and C. papyrus (Visser, 1964), although not for Typha latifolia (Boyd, 1970). This does not preclude a net loss of nitrogen from decaying litters, and such release has been demonstrated for Scirpus fluviatilis and T. glauca (Davis and van der Valk, 1978b). Nitrogen dynamics may in fact vary under different experimental conditions of aeration and temperature (Godshalk and Wetzel, 1978a). Andersen (1978) reported that nitrogen levels increase in P. australis litter in a eutrophic lake, but not in a nutrient- poor one. Decomposition rate may correlate positively with initial nitrogen content (Davis and van der Valk, 1978a). Much of the nitrogen immobilis- ation may be due to microbial uptake, and some of the increase in its levels may be in the form of proteins; complexing with lignins has been suggested for decomposing tree leaves (Suberkropp et al., 1976b).

Phosphorus may also exhibit a rapid initial fall in level (Boyd, 1970; Hod- kinson, 1975a; Mason and Bryant, 1975a; Chamie and Richardson, 1978). Much of this released phosphorus, however, should be reabsorbed (Planter, 1970; Howard-Williams and Howard-Williams, 1978), and some litters may show a net uptake of this element (Davis and van der Valk, 1978b).

Even where the nitrogen level increases significantly, the carbon:nitrogen ratio of the litter may still remain high (Hodkinson, 1975a; Best et al., 1982). This ratio is usually far greater than the 17: 1 value proposed as critical for animal nutrition (Russell-Hunter, 1970) and the 1O:l level suggested as the optimal one for microbial decomposition by Alexander (1971). Although some soluble phenolic substances may be lost early (Suberkropp et al., 1976b; Howard-Williams and Junk, 1977; Howard-Williams and Howard- Williams, 1978), other refractory constituents commonly increase in the litter (Visser, 1964; Polunin, 1982b). The dynamics of structural substances such as lignin, cellulose, and hemicellulose can be influenced by environmental conditions (Godshalk and Wetzel, 1978a). The net effect of all these changes is probably to reduce susceptibility to decomposition, as shown for P. australis (Fig. 3).

Howard-Williams and Junk (1977) found that with time the variance of chemical constituent concentrations of decomposing Amazonian floating-


201 o ~ . i ~ ~ ; ~ ~ b o IeOlign,in , o

10 - - I

0 50 150 250 350 2 0

- s D

Nitrogen 2 1 s%” o o o o o

I o,6t !:ic d;bility

0

0 1

I

.- m

0 ‘ . 0 a,

0 50 150 2M 350

0.4

0.2

0 0 0 0 0 n . n

0 50 150 250 350 Time (days)

Fig. 3. Changes with time (in days) in the concentration of refractory components (as acid- insoluble lignin), nitrogen (Kjeldahl), and organic decayability (mean ash-free weight loss per day of particles > 61 < 243 p n at 16°C in aerated bags in the laboratory between days 12 and 30) of Phrugrnites uustrufis leaf litter exposed as loosely tied bundles in the shallow littoral of the main brick pond, Wicken Fen, Cambridgeshire, England. After Polunin (1982b).

meadow grasses declined. In some situations therefore, environmental factors may have a dominant influence on the final constitution of the detritus.

VII. DECOMPOSITION OF DISSOLVED AND PARTICULATE MATTER

Much of the organic matter produced by emergent macrophytes sooner or later assumes dissolved, colloidal, or particulate detrital forms. Under experimental conditions, the fate of the litter organic matter is influenced greatly by temperature and aeration: in a warm aerobic environment, most of the matter is probably reduced to carbon dioxide, but in cold anaerobic conditions accumulation in sediments is likely to predominate (Fig. 4).

It is convenient here to split these phases into those which are, at least temporarily, in the water column and those which are deposited on the bottom. In practice, dissolved matter is defined as that which passes through a filter with a pore size of about 0.5 pm. Particulate matter, whether suspended or deposited, is anything larger than this. The “dissolved” fraction may also contain some colloidal matter.


sediment I sediment

naerobic Aerobic lXJTFI.ls.cl I

<1 e’c 43’c Anaerobic Aerobic

co2

R D C M t LWM RDOM f L W M

sediment sediment

Fig. 4. Effect of temperature variation and anaerobic vs aerobic conditions on macrophyte detritus in laboratory microcosms. The proportion of detritus reduced to each of four states is represented, where “CO,” indicates organic matter completely mineralised and “sediment” is matter which was not decomposed over some 180 days and accumulated as particulate deposits; “RDOM” and “LDOM” are refractory and labile soluble organic matter, respectively. After Godshalk and Wetzel (1977).

A. Suspended and Dissolved Organic Matter Some dissolved organic matter may be derived from live emergent ma-

crophytes (Szczepadska and Szczepanski, 1973). A major source, however, is the leachate from freshly inundated detritus (OlAh, 1972; Otsuki and Wetzel, 1973; Howard-Williams and Howard-Williams, 1978). Later on in decomposition, dissolved and colloidal organic matter may also accumulate from microbial activity and macroinvertebrate feeding (Polunin, 1982b). This release occurs under a wide range of conditions of temperature and aeration, but produces distinct fractions of dissolved matter which are more or less labile (Godshalk and Wetzel, 1977). Inputs of dissolved and colloidal matter may be gained by waters passing through littoral marshes into lakes and rivers (Wetzel and Otsuki, 1973).

When litter input is high, as on senescence at higher latitudes, an increase in dissolved organic matter levels in the receiving waters is expected. This is not necessarily found, even in small water bodies lined with reeds, however, and the dissolved matter must either be utilised quickly or be con-

140 NICHOLAS V..C. POLUNIN

verted to some other form (Polunin, 1982a). Bacteria seem to be particularly important in the decomposition of such material (OlAh, 1972; Wetzel and Manny, 1972), and large numbers of these may be present in the littoral water column associated with macrophyte decomposition (Krashennini- kova, 1958; Olkh, 1969a,b, and Aliverdieva-Gamidova, 1969, both reported in Wetzel, 1975, p. 583). Some bacteria are also able to use humic materials (de Haan, 1972, 1974). Fungi do not appear to be significant, at least in the early phases of decomposition of this material (OlAh, 1972; Wetzel and Manny, 1972). Much of the labile organic matter may thus be assimilated by bacteria and enter a particulate phase as living microbial biomass (cf. Robertson et al., 1982). In hard water much of the dissolved matter may be absorbed onto carbonate particles and precipitated in this form (Wetzel, 1969; Wetzel and Manny, 1972; Otsuki and Wetzel, 1973), and particle formation from it can be enhanced by turbulence (Lush and Hynes, 1973). Inorganic particles may also accelerate bacterial utilisation of dissolved matter (Paerl, 1977). Bacteria may ultimately be ingested by metazoan consumers such as cladocerans (Peterson et al., 1978), and cal- anoid copepods, crustaceans, sphaeriid bivalves, chironomid larvae (Mon- akov, 1972), gastropod molluscs (e.g., Mason, 1977, p. 41), and the aquatic toad Xenopus laevis (McConnell, 1968).

Suspended fine particulate matter, whether directly from litter decomposition or secondarily derived from dissolved matter, may be used by some zooplankton (Saunders, 1972; Monakov, 1972) and benthic filter feeders (Wallace and Merritt, 1980) in particular. Such suspended particulate matter may sometimes be derived from littoral deposits (Moss, 1970; Polunh, 1982a), and emergent macrophytes, where they are present, typically make a substantial contribution to this material (Wetzel, 1975, p. 586). The suspension can be caused by water movement induced by wind or differential heating of littoral waters (Schroder, 1975; Saunders et al., 1980), and also by the feeding activities of fishes such as carp (Wetzel, 1975, p. 530). Benthic detritivores which break down the litter to finer particles (Mason and Bryant, 1975a; Polunin, 1982b; Kirby et al., 1984) can be expected to increase the probability of resuspension.

Export of organic matter from littoral swamps may be particularly significant in water bodies where such resuspension is effective and the ratio of the littoral to the open water area is large. Where export of organic matter does occur, the rate of its decomposition, and fate overall, remain unknown. Hargrave (1975) suggests that for stratified lakes the depth of mixing may be important in determining how much of the suspended particulate matter is decomposed before it is deposited. Whether this would be so for most organic matter from emergent stands is unsure, because this


material consists typically of degraded components rather resistant to decay.

B. Deposits The particulate material accumulating in emergent macrophyte swamps

may come from many sources. Allochthonous matter may be important in some cases, especially that from flood plains, but such material is often largely inorganic (Walker, 1970). Littoral plants such as emergents may locally enhance the accumulation by increasing deposition and reducing erosion (Richards, 1934; Buttery et al., 1965; Fisher and Carpenter, 1976; Alizai and McManus, 1980). Autochthonous organic matter, however, can be substantial in amounts; more than three-quarters of the freshwater strati- graphic data on such succession reviewed by Walker (1970) suggested accumulation rates between 21 and 80 cm per thousand years. On high tidal marsh dominated by Phragmites an annual rate of 17.1 mm has been estimated (Harrison and Bloom, 1977). In such cases the influence of water movement may be considerable, while in lakes wave action (Pieczyhka, 1975) and ice (Uotila, 1971) may mechanically break down emergent litter and help to deposit it on the shore.

Emergent macrophytes make substantial contributions to limnic (below water level, e.g., Carex rostrata) and telmatic (periodically inundated, e.g., Phragmites australis, Cladium mariscus) peats (Moore and Bellamy, 1974). Godwin (1975) has remarked on the frequency of occurrence of identifiable fragments of P. australis in sub-fossil deposits, in contrast to other Gra- mineae. Most deposited material, however, is in particulate form (Walker, 1970). Such accumulation is enhanced by low temperature, pH, and wa- terlogging. Although the carbon:nitrogen ratio may fall with increasing depth into the peat of papyrus swamps, the contribution made to the re- maining matter by humic compounds, and also acidity, increase (Visser, 1964). For such reasons the decomposition rate typically declines at greater depths (Moore and Bellamy, 1974). Particularly where emergent plants form a floating mat of vegetation on the surface of the water, anaerobic conditions can be expected to reduce decomposition rate and enhance accumulation underneath; such deposits evidently constitute an important sink for certain inorganic nutrients, at least temporarily (Moore and Bellamy, 1974; Gaudet, 1976).

Swamp soils high on the littoral are commonly exposed to a seasonal alternation of conditions. Rewetting of littoral sediments frequently yields a large, short-lived flush of certain ions such as phosphate (Moss and Moss, 1969; Howard-Williams, 1972). Decomposition is inhibited during periods


of complete drying, but on rewetting a burst of microbial activity may occur (Stevenson, 1956; van Schreven, 1967; Sorensen, 1974; but see Bryant et al., 1982). Carbon and nitrogen release can be greater on rewetting the longer the period of desiccation, and soils high in humic substances may yield relatively large amounts of nitrogen (Birch, 1960). Alternate periods under aerobic and anaerobic conditions may also be beneficial to decomposition, although the rate in a constantly anaerobic environment is less than that in a constantly aerobic one; the higher the frequency of alternation, the greater can be the release of total nitrogen (Reddy and Patrick, 1975).

At least in fairly new material, inorganic nutrients such as available forms of nitrogen and phosphorus may speed decomposition (Polunin, 1982b), and particle size may also play an important role, decomposition rate being inversely related to it (Hargrave, 1972). In the cases of both of these factors, however, there is evidence that the enhancement may affect only part of the detrital resource (Polunin, 1982b). Supplementing nitrogen and phosphorus in the medium does not necessarily stimulate decomposition of refractory components such as lignin (Waksman and Tenney, 1928; Federle and Vestal, 1980b). This is significant because lignin is an important part of freshwater sediments (Vallentyne, 1957; Farmer and Morrison, 1964). Much natural detritus from lentic waters thus has the characteristics of highly degraded material in which bacterial numbers Nlehlovd, 1978a,b; Oldh, 1972) and oxygen uptake (Hargrave, 1972) are much reduced. Such aged material is also accepted by detritivores less readily than detritus freshly derived from leaf litter (Ward and Cummins, 1979).

Nevertheless, large seasonal changes in swamp sediments at high latitudes can be expected (Dokulil, 1975). Rates of methane evolution from deposits may increase with rising temperature (Baker-Blocker et al., 1977; Lovley and Klug, 1982). Most of this activity, however, is confined to the upper layers of sediment (Hayes and Anthony, 1959; Harrison et al., 1971; Hob- bie et al., 1980). Local additions of inorganic nutrients (Hobbie et al., 1980) and of organic matter (Kelly and Chynoweth, 1981) can stimulate microbial activity. These microbial processes in the sediments are an important component of the swamp ecosystem, because microbial activity is concentrated there (Dokulil, 1975).

Deposits derived from emergent macrophytes generally constitute a favourable habitat for aquatic animals (McLachlan, 1974; Pieczynska, 1977), and detritivores living in or on the sediment may in turn influence the microbial processes. Species such as the oligochaete Limnodrilus (Monakov, 1972) and the gammarid Hyalella (Hargrave, 1970b) ingest deposits apparently without discrimination. These and other detritivores such as the simuliid and chironomid larvae do, however, selectively digest microor-


ganisms such as bacteria (Baker and Bradnam, 1976). As a consequence, most of the material ingested is not assimilated and passes through the gut very quickly: simuliid larvae (Ladle et af., 1972) and Hyafefla (Hargrave, 1970b) pass ingested matter through in 30 minutes or less, while Limnod- rifus ingests two to eight times its own body weight in deposit daily (Mon- akov, 1972). Hyalelfa, however, assimilates surface sediment up to 0.5 mm depth more efficiently than subsurface sediment at a depth of 2 cm; even over such small distances, therefore, the nutritional suitability probably declines markedly.

Detritivores commonly stimulate microbial activity and also decomposition of deposited material. Burrow-forming species such as many oligochaetes (Edwards, 1958; Davis, 1974) and chironomid larvae (Hobbie et al., 1980) have been shown to be able to modify redox profiles of the sediment, in particular increasing the depth of the oxygenated layer. Chironomids may thus greatly enhance the total numbers of bacteria (Kajak et al., 1968) and overall oxygen consumption by the sediment (Rybak, 1969), although the reverse, a decrease in respiration, has been found by Hobbie et al. (1980) at low temperature. Under similar conditions in an Arctic tundra pond, however, the tadpole shrimp Lepidurus led to an increase in microbial respiration and also to faster decomposition of plant litter (Hobbie et al., 1980). At natural densities Hyalella stimulates bacterial production in lake sediment, as well as primary productivity and community respiration as a whole (Hargrave, 1970a). Microbial decomposer enhancement in this case may come about through mechanisms such as the release of dissolved organic matter, or increasing particle surface area; like the burrowing species, Hy- alefla may serve to aerate the substratum. Evidence for aeration of deeper sediment by the roots and rhizomes of emergent macrophytes has been presented (Andersen and Hansen, 1982).

VIII. THE SWAMP ENVIRONMENT

An account has been given above of the main features of the detrital system based upon emergent macrophytes. This description will now be rounded off by considering some distinctive ecological features of these swamps, particularly as they affect, and in turn are influenced by, decomposition.

A. Physicochemical Environment Emergent swamps exhibit considerable spatial variability in their envi-

ronment (Planter, 1973; Ulehlovd et al., 1973; Howard-Williams and Len- ton, 1975; Howard-Williams, 1979b). This is particularly noticeable in


vertical transects from land through an emergent macrophyte stand to open water (Fig. 5). The swamp environment itself is also often seen to be quite distinctive, in addition varying within itself (Fig. 6). The pH and dissolved oxygen content of the swamp water are typically low (Carter and Beadle, 1930; Carter, 1934; Gorham, 1953; DvofBk, 1970a; Welcomme, 1970; Howard-Williams, 1972; Pieczydska, 1973, 1975; Rzbska, 1974; Gaudet, 1977; Howard-Williams, 1979a). The acidity is influenced by the production of carbon dioxide and the release of hydrogen ions from some peats. Oxygen depletion is enhanced by decomposition and probably some abiotic uptake, and also by reduced algal photosynthesis and wind-induced mixing. Turbidity can fall inside emergent macrophyte stands, and levels of total dissolved organic matter and conductivity may be higher there (Banoub, 1975; Howard-Williams, 1979a).

Deposited material derived from emergent macrophytes may constitute an important reservoir of limiting mineral nutrients. In this, interstitial water can be particularly important for concentrations of ions such as ammonium, phosphate, and iron tend to be many times greater there than in the overlying littoral water (Planter, 1973). These high concentrations evidently depend on a number of factors, including the rate of mineralisation from plant detritus, alkalinity, redox potential, plant colonisation, and inflow from drainage. Particularly under anaerobic conditions, chemical release of phosphate (Mortimer, 1941) and other substances such as ammonia (Kamp-Nielsen, 1974) may occur. The particulate matter accumulating under emergent macrophyte stands may also be important in nutrient path-

9

10 8 6 4 2 0 Distance into swamp from lake (km)

7

Fig. 5. Change in the conductivity and pH of soil and water along an 11-km transect through the northern swamp of Lake Chilwa, Malawi. After Howard-Williams (1979a).

DECOMPOSITION OF MACROPHYTES

200-

E

E .ki 100-

0 al I .-

0 -

145

swamp i pond

r-' - 8

- / O 7 15

P -0- A0

concentration

200

0-0 Phytoplankton 0 photosynthesis

5 Respiration

I€ P A 0 C D E A

Fig. 6. (above) Diagram of a transect through a stand of Glyceria aquatica on the littoral of Radov Pond, southwestern Bohemia, Czechoslovakia, showing the position of five main stations at which water samples were taken for determination of (below) mean temperature, alkalinity, colour, community respiration, pH, dissolved oxygen concentration, and phytoplankton photosynthesis in July to September of 1965. After Dvoidk (1970a).

ways: in the Typha glauca stand studied by Davis and van der Valk (1984) there was a net accumulation of calcium and nitrogen, and a loss of sodium and potassium.

Significant nitrogen fixation can take place in emergent macrophyte sediments; under Glyceria borealis, for example, nitrogen is fixed at a rate


equivalent to half of the plant’s requirement of this element for growth (Bristow, 1975). Where water flows seasonally through swamps, considerable amounts of certain elements may be exported, at least temporarily, to open water (Howard-Williams, 1972, 1979a; Gaudet, 1979). Such release from littoral swamps also occurs continually at the interface of the vegetation with the open water, where water mixing may be induced by wind (Howard-Williams, 1979a) or convection (Schroder, 1975). There is also evidence that significant nutrient release occurs from the deposits below floating emergent swamps where the sediments are exposed to aerobic conditions (Gaudet, 1976). Such releases may serve to explain the large increase in primary productivity which can occur at the open-water margin of swamps (Rodewald-Rudescu, 1974; Gaudet, 1976; Howard-Williams, 1979a). Although deoxygenation may be unfavourable to most organisms, nevertheless anaerobic benthic microbial processes will commonly make a substantial contribution to carbon (cf. Rich and Wetzel, 1978) and nutrient (cf. Wiebe, 1979) flows. The proximity of aerobic and anaerobic phases in the ecosystem may help to explain the high productivity of reed swamps.

B. Succession Emergent macrophytes are important contributors to the hydrosere, the

biotic succession from water to land, at high latitudes (Tansley, 1939; Con- way, 1949) and in the tropics (Flenley, 1979). British stratigraphical data suggest that reed swamps have occurred as a stage in some three-quarters of autogenic successions which have been analysed, and that this stage may be a crucial element in large bodies of water in particular (Walker, 1970). Because of high autogenic matter accumulation the stage is usually short- lived: succession from a swamp with floating-leaved macrophytes to a fen, which has no permanent standing water, commonly takes less than lo00 years. This accretion occurs because even though much of the litter is broken down, a significant part of the breakdown products accumulate in situ. Both environmental conditions, such as low temperature and oxygen de- ficiency, and factors intrinsic to the detrital material, such as its high content of structural fibres and high carbon:nitrogen ratio, are conducive to this inefficiency of decomposition, the failure to mineralise the detrital matter completely. The plants themselves therefore contribute to this state of affairs in several ways, and the poor turnover of carbon by decomposition is so prominent a feature of swamps (Given, 1975) that one wonders whether it has not been a factor in the success of emergent macrophytes, which specialise in these marginal environments.

Accretion may be accelerated by allogenic sedimentation, which is im-


portant in many cases (Spence, 1967, 1982), and which is also enhanced by the plants themselves (Buttery et al., 1965; Alizai and McManus, 1980).

C. Large Aquatic Consumers Voigts’ (1976) study of northern Iowa marshes demonstrated qualitative

and quantitative differences in macroinvertebrate assemblages characteristic of various types of vegetation. Organisms associated with emergent macrophyte swamps presumably also come and go in succession, although such data do not appear to have been presented yet.

The organisms, like the abiotic environment, can be highly variable, as indicated by the study of Macan and Kitching (1972) on macroinvertebrate abundance and species composition. This heterogeneity exists especially along vertical transcents perpendicular to the shore. Densities, assemblages, and total biomass of species can change greatly across the swamp (Mesch- kat, 1934; DvofAk, 1970a; Mason and Bryant, 1974; Fig. l), and in the open water beyond (Kajak and Dusoge, 1975a,b).

Animals living inside swamps have to be able to cope with conditions such as the low pH and oxygen depletion which commonly prevail there. Invertebrates such as leeches (Mann, 1956) and fishes (Fish, 1956), for example, may have unusual respiratory physiologies; many species are air breathers (Dvoidk, 1970a; Beadle, 1974). There are few submerged macrophytes in these swamps, except in clearings (Carter and Beadle, 1930; Howard-Williams, 1972), and algal production tends to be greatly reduced DvoiAk, 1970a; Dokulil, 1973; Howard-Williams, 1979b; Fig. 6). The emergent macrophytes themselves therefore make a substantial contribution to the organic production of many freshwater systems (Ambasht, 1971; Barko et al., 1977; Howard-Williams, 1979b; Howard-Williams and Junk, 1977; Wetzel and Hough, 1973; Komkkovd and Komdrek, 1975). The view has often been expressed (e.g., Saunders, 1972; Hill and Webster, 1982) that plant litter, particularly its gradual decomposition, may trophically sta- bilise freshwater communities. The evidence for this idea, however, is slight, particularly in lentic systems.

Emergent macrophyte material, however, is trophically significant in many fresh waters. While the original living plants themselves are often not particularly acceptable to many consumers, the particulate detritus is, at least in some respects, more likely to be so. The peat and sludge below Cyperus papyrus swamps are, for example, enriched in nitrogen (Gaudet, 1976). The detrital substrate usually becomes increasingly refractory to decomposition in the long term, but detritivores are evidently effective at di- gesting microorganisms on such material. If they can rapidly process the particulate material, assimilating the useful components efficiently, then


the detritus may be a reusable source of food over a long time period. Nevertheless, if fresh material is available it will typically be preferred, and may be consumed more rapidly (Kdodziejczyk and Martynuska, 1980). Such fresh material may be provided seasonally for detritivores at higher latitudes, and more uniformly through the year for those at lower latitudes.

The particulate matter produced in abundance by the decomposition of emergent macrophytes may be utilised by larger animals such as fishes directly or indirectly. Trophic opportunism appears to be a common characteristic of many macrophyte-dominated systems (Berrie, 1976; Kikuchi and Ptrbs, 1977; Welcomme, 1979). This may be a feature in particular of seasonality, along with the tendency to take items other than plant detritus, when these are available (e.g., Mason and Bryant, 1975b; Smork and Stone- burner, 1980) at higher latitudes; this is where most relevant work has been carried out.

Specialised deposit feeders, however, are prominently associated with tropical swamps in particular. Many mud-eating fishes have been men- tioned in the literature, for example from the Amazon (Marlier, 1967), the Congo (Marlier, 1958), and the Nile (Sandon and el Thayib, 1953). There are many species which ingest coarser detrital material predimenantly, in major river systems of South America (Marlier, 1967) and of central Africa (Matthes, 1964). A problem with all such detrital feeding, however, is to know where the detritus comes from, and what its nutritive role is; little work has been carried out on either topic.

Fishes may also be dependent on detrital systems of swamps in a more indirect way. This dependency can be through feeding on detritivorous invertebrates (e.g., Chambers, 1971; Prus, 1977), which may be more abundant around emergent swamps (Pieczyhska, 1977; Moss, 1979). It may also be through consuming planktonic organisms, the production of which can be enhanced (Rodewald-Rudescu, 1974), possibly by the release of primary nutrients at swamp margins (Gaudet, 1976) or, alternatively, by the availability of detrital organic matter and the decomposer microorganisms which utilise it. Cladocerans, for example, can be abundant at the open- water edge of emergent swamps (Thomas, 1961; Rzbska, 1974), and there is evidence that density, and also species diversity, may be controlled by the local efficiency of decomposition (Smyly, 1955). Some of these species ingest bacteria and minute detrital particles, while others are phytoplankton grazers (Monakov, 1972).

D. Potential Economic Importance Whatever the precise reasons, littoral swamps are widely regarded as a

significant habitat in fisheries, whether in lakes (Harding, 1960; Beadle, 1974; Gwahaba, 1975; Howard-Williams and Lenton, 1975; Howard-


Williams, 1979b) or in river systems (Stubbs, 1949; Daget, 1952; Marlier, 1958; BacalbaSa-Dobrovici, 1971; Rzdska, 1974). In both types of environment, the detrital systems to which emergent macrophytes contribute can be expected often to be important. Mud-eating fishes of the genera Cith- arhinus (Marlier, 1958), Labeo (Sandon and el Thayib, 1953), and Pro- chilodus (Bonetto et al., 1969) are of widespread economic value, for example. Seasonal flooding is a chief characteristic of swamps (Talling, 1957; Mizuno and Mori, 1970; Beadle, 1974), and many fishes depend on the inundation of littoral areas to feed and breed (Daget, 1952; Williams, 1972; Rzbska, 1974).

The economic values of decomposition processes cannot be estimated as yet; they are only indicated by circumstantial evidence. Until quite recently the view was expressed more than once that emergent swamps should be cleared to increase fishery yields, as the anoxic conditions were assumed to be unfavourable (e.g., Hickling, 1961). It is now appreciated, however, that emergent macrophyte swamps may often provide a valuable basis for high natural organic production and nutrient cycling, and their margins can be zones of intense production of zooplankton and fish. Regrettably information on this relationship in mechanistic-ecological and economic terms remains poor. Szajnowski (1970), for example, found a positive correlation between the yield of tench and predatory pike in lakes and the proportion of the lake area occupied by the littoral, but there was no correlation specifically with the production of reed. Krzywosz et al. (1980) observed, however, that when beds of Phragmites and Typha were severely damaged by the grass carp Ctenopharyngodon idella in a Polish lake, many fish.species declined in abundance; they proposed that this was due to loss of shelter previously afforded by the plants. Williams (1972) has reported a positive correlation between water level and fishery yields in two Zambian lakes; high-water years inundate littoral marshes where the adults can breed and the young may be better protected from predation. Thompson (1977) men- tions that in Lake Huleh in Israel, a papyrus swamp was cleared in the hope that the original plant production would be matched by future fish culture; it happened that it was not, for nutrients were lost with the vegetation and with the erosion of the swamp soils. Gorbunov (1953) felt that, through their detritus, emergent macrophytes were a major source of organic matter in some Russian rivers, and that the detritus exported by spring floods must ultimately be essential to certain fisheries. Such a role has been indicated for young chum salmon, Oncorhynchus keta, which feed on harpacticoid copepods associated with the detritus (Sibert et al., 1977). As often as they may be attributable to a direct trophic role, therefore,

the fisheries functions of emergent macrophyte swamps are explicable in terms of shelter and the provision of a substrate for high local periphyton production, as suggested by examples reviewed in Lowe-McConnell(l975)


and Pieczyriska and Ozimek (1 976). Decomposition probably contributes indirectly to the latter, however, because a high level of mineralisation en- hances epiphyte production through the release of nutrients. Knowledge of the manner in which emergent macrophytes may contribute to fisheries is thus at an elementary stage.

Apart from potential functions in fisheries, production and decomposition processes in emergent macrophyte swamps could also be of value in the small-scale treatment of sewage. Studies in Holland have shown, for example, that ponds dominated by Scirpus lacustris and Phragmites australis can substantially reduce the loads of suspended matter, nitrogen, and phosphorus in sewage effluent (de Jong and Kok, 1978). For local appli- cations such ponds could be effective, and also cheaper to run than activated-sludge plants.

One must therefore agree with Moore (1980) when he concludes that the clearing of these swamps on a large scale is unwise before their value is better understood. A major objective of future research should thus be to investigate trophic and other roles of swamps which are of economic benefit to man.

The plight of these habitats epitomises that of many ecotones, caught as they are in a no-man’s land between two biomes. In research terms it appears that limnologists have too often viewed these marginal areas as being primarily the domain of the terrestrial scientist; the terrestrial specialist has in turn regarded them as being too aquatic for his own interests. As far as exploitation is concerned, aquatic and land users of swamps too often take benefits which they reap for granted, without assuming the responsibility for those gains. In addition, adverse effects of swamps on man are be- moaned, while uses of these areas are commonly in conflict with each other. Such incompatibilities will not be sustainable for much longer in many regions.

IX. SUMMARY: FATE OF THE DETRITUS

Although outbreaks of grazing animals occur from time to time, and factors such as fire may occasionally dissipate a large proportion of the high net primary production, most of the large amount of organic matter generated by emergent macrophytes experiences death undisturbed and thus enters the detrital pool directly. The litter input varies in quality and amount both in time and space; the temporal changes depend particularly on the life histories of plants, and the spatial differences are influenced by geo- graphical changes within species and in the composition of communities. These sources of variation are related on a large scale to environmental conditions.


To begin with, free-fallen litter may float on the water and, particularly at the margin of the reed swamp, it can be transported away from the site where it landed. The early stages of decomposition in fresh water are dominated by leaching, although the role of leaching-type processes later on has not been assessed. Leaching weight loss may be affected by environmental conditions to a small degree, but its extent is a function primarily of the amount of soluble matter present. Early leaching makes a significant contribution to weight loss, amounting commonly to more than 10% of the leaf litter. Organic constituents make a significant, but an as yet unmea- sured, contribution to this leaching loss.

Microbial colonisation is rapid and a succession of forms may occur on the litter, but these early colonisers, mostly bacteria, are apparently not actively contributing to decomposition of the detritus and are merely using components of the soluble leachate. Benthic macroinvertebrates increase in total numbers and variety with time, but much of this colonisation appears unaffected by changes in the litter itself.

Comminution may begin with abiotic processes such as ice-scour and wind acting on dead-standing plants. Macroinvertebrate breakdown commonly accelerates after a lag phase. Although in energetic terms these detritivores may not be significant, it is evident that where they are abundant a great deal of the litter can be comminuted by them, even though it may not be nutritionally valuable. The particulate matter resulting from breakdown is a major part of the organic accumulation within the swamp.

Microbial decomposition is a dominant process in litter decomposition; it can be influenced especially by environmental conditions (temperature, oxygen and nutrient availability, moisture) and resource quality (inhibitors, refractory components, intrinsic nutrients). There is little specific information on the microorganisms which carry out this activity. Animals can influence microbial processes in a number of ways, and may affect decomposition indirectly.

Two major pathways, accumulation and mineralisation, play a dominant role in accounting for the detrital input within reed swamps. Export away from littoral reed swamps could also be locally significant; the extent of this depends largely on particle size and water movement. The balance between accumulation and mineralisation is influenced both by environmental factors and by intrinsic characteristics of the detritus. Both types of factor may interact positively (inorganic nutrients available from the water may accelerate mineralisation just as those available in the detritus itself may do) or negatively (the accumulation of lignin in the detritus may offset the beneficial effect of an inorganic nutrient supply in the water).

It is not yet possible to produce a detrital organic matter budget for any emergent macrophyte. The most detailed work to date is that on species such as Scirpus acutus, Phragmites australis, and Carex aquatilis. In order


to produce quantitative information on the fate of the detritus it is necessary to consider not only how the input of detrital matter varies through the year, but also how the environment changes seasonally in factors influential to decomposition, and how mineralisation of organic matter re- sponds to both processes. This type of work has not yet been carried out.

Although methods such as litter bags employed to assess decomposition are crude, results from them would be more valuable if the experimental conditions were reported in greater detail than they often are (Larsen, 1982). Both quality of the material and the conditions under which it decays are so influential in decomposition that more attention should be given to de- tails of the litter and environment being studied. The manner in which the material is prepared and presented warrants greater care than it is commonly now accorded. Important points here are the timing and density of the experimental input relative to the natural input. Another influential factor is the way in which experimental litter is dried. Drying at 100-150°C can alter both the rate of decomposition, either increasing (Hobbie et al., 1980) or decreasing it (Hofsten and Edberg, 1972), and its characteristics (Harrison and Mann, 1975). Evidently then, the use of such high temperatures is unwise, but lower temperatures can also modify decomposition. Unless controls are incorporated, it may thus be better to avoid oven-drying altogether.

Laboratory studies have begun to provide useful information on decomposition processes and breakdown products under specified conditions, but rather little is known about how such experimental findings relate to the field.

For detailed budgets, information is needed on the early leachate and on how rapidly dissolved organics are utilised biologically or converted into other forms. Similarly, the fate of particulate and dissolved matter derived from decomposition later on needs to be better known. How rapidly is the material produced at different times of the year? How quickly is it then decomposed? How likely is export'of such material to occur? What proportion of the detrital input tends to accumulate in situ, and what factors most determine this accretion? The linkage between short-term estimates of litter breakdown and measurement of the rate of accretion in the long term remains an important gap to be filled. It is essential for such work to know how long it effectively takes for the detritus derived from any one year's production to reach a refractory state, and also what proportion of the whole input that remnant represents. With such data it could then be possible to see how estimates of accumulation compared with measurements of accretion from stratigraphical work.

Although the turnover of above-ground parts through decomposition may occur at a greater rate and be of greater significance to food chains and the


recycling of mineral nutrients than that of below-ground parts, it is clear that far more information is needed on the latter than is at present available. In certain respects, its spatial uniformity for example, the underground environment may be simpler than the above-ground one, but problems of assessing the detrital input and decomposition of rhizomes and roots would seem to be greater than for aerial parts. It may thus be a long while yet before this aspect, vital though it is to an understanding of the fate of organic production, is adequately studied.

ACKNOWLEDGEMENTS

I am grateful to A. Macfadyen, R. G. Wetzel, B. Moss, C. Howard-Williams, and C. F. Mason for their useful comments on the draft of this article. I express my appreciation to C. Howard-Williams and R. G. Wetzel again, and also to J. DvoHk, for permitting me to adapt previously published data of theirs to the present account. Finally I thank Professor 0. Horn for providing me with facilities in his department.

REFERENCES

Adcock, J. A. (1982). Energetics of a population of Asellus aquaticus (Crustacea, Isopoda): respiration and energy budgets. Freshwater BioI. 12, 259-269.

Alexander, M. (1971). “Microbial Ecology.” Wiley, New York. Alexander, V., Stanley, D. W., Daley, R. J., and McRoy, C.P. (1980). Primary producers.

In “Limnology of Tundra Ponds” (J. E. Hobbie, ed.), pp. 179-194. Dowden, Hutchinson & Ross, Stroudsburg. Pennsylvania.

Aliverdieva-Gamidova, L. A. (1969). Microbiological processes in Lake Melchteb. Mikro- biologiya 36, 1096-1 100.

Alizai, S. A. K., and McManus, J. (1980). The significance of reedbeds on siltation in the Tay Estuary. Proc. R. SOC. Edinburgh, Sect. B: BioI. 78, 500-515.

Ambasht, R. S. (1971). Ecosystem study of a tropical pond in relation to primary production of different vegetational zones. Hidrobiologia 12, 57-61.

Andersen, F.O. (1978). Effects of nutrient level on the decomposition of Phragmitescommunis Trin. Arch. Hydrobiol. 84, 42-54.

Andersen, F. O., and Hansen, J. I. (1982). Nitrogen cycling and microbial decomposition in sediments with Phragmites australis (Poaceae). Hydrobiol. Bull. 16, 11-19.

Anderson, N. H., and Grafius, E. (1975). Utilization and processing of allochthonous material by stream Trichoptera. Verh. Int. Ver. Theor. Angew. Limnol. 19, 3083-3088.

Anderson, A., and Danell, K. (1982). Response of freshwater macroinvertebrates to addition of terrestrial plant litter. J. Appl. Ecol. 19, 319-326.

Aphis, A. E., Chesters, C. G. C., and Taligoola, H. K. (1972). Microfungi colonizing submerged standing culms of Phragmites communb Trin. Nova Hedwigia 23, 473-480.

Assawah, M. W. (1956). Root-region fungi of Iris and other plants. Ph.D Thesis, University of Nottingham.

BacalbaSa-Dobrovici, N. (1971). The economic importance of the burgu (Echinochloa stagnina P. Beauv.) on the Middle Niger in the Niger Republic. Hidrobiologia 12, 41-46.


Baker, J. H., and Bradnam, L. A. (1976). The role of bacteria in the nutrition of aquatic detritivores. Oecologia 24, 95-104.

Baker-Blocker, A., Donahue, T. M., and Mancy, K. H. (1977). Methane flux from wetland areas. Tellus 29, 245-250.

Banoub, M. W. (1975). The effects of reeds on the water chemistry of Gnadensee (Bodensee). Arch. Hydrobiol. 75, 500-521.

Barko, J. H., Murphy, P. G., and Wetzel, R. G. (1977). An investigation of primary production and ecosystem metabolism in a Lake Michigan dune pond. Arch. Hydrobiol. 81,

Barlocher, F., and Kendrick, B. (1975a). Leaf conditioning by microorganisms. Oecologia 20,

Barlocher, F.. and Kendrick, B. (1975b). Assimilation efficiency of Gammarus pseudolim- n a m (Amphipoda) feeding on fungal mycelium or autumn-shed leaves. Oikos 26, 55-59.

Barlocher, F., and Kendrick, B. (1976). Hyphomycetes as intermediaries of energy flow in streams. In “Recent Advances in Aquatic Mycology” (E. B. Gareth Jones, ed.), pp. 435- 446. Elek Science, London.

Barsdate, R. J., Prentki, R. T., and Fenchel. T. (1974). Phosphorus cycle of model ecosystems: significance for decomposer food-chains and effect of bacterial grazers. Oikos 25,

Bayly, I. L., and O’Neill, T. A. (1972). Seasonal ionic fluctuations in a Typha glauca com-

Beadle, L. C. (1974). “The Inland Waters of Tropical Africa: an Introduction to Tropical

Bernard, J. M., and Solsky, B. A. (1977). Nutrient cycling in a Carex lacustris wetland. Can.

Berrie, A. D. (1976). Detritus, micro-organisms and animals in freshwater. In “The Role of Terrestrial and Aquatic Organisms in Decomposition Processes” (J. M. Anderson and A. Macfadyen, eds.), pp. 323-338. Blackwell, Oxford.

Best, E. P. H., Zippin, M., and Dassen, J. H. A. (1982). Studies on decomposition of Phrag- mites australis leaves under laboratory conditions. Hydrobiol. Bull. 16, 21-33.

Birch, H. F. (1960). Nitrification in soils after different periods of dryness. Plant Soil 12,

Bjamov, N. (1972). Carbohydrases in Chironomus, Gammarus and some Trichoptera. Oikos

Blake, G. (1982). Characterization of decomposition of plant material in an Alpine lake. Hy- drobiol. Bull. 16, 5-9.

Bobbie, R. J.. Morrison, S. J., and White, D. C. (1978). Effects of substrate biodegradability on the mass and activity of associated estuarine microbiota. Appl. Environ. Microbiol. 35,

Bonetto, A., Dioni, W., and Pignalberi, C. (1969). Limnological investigations on biotic communities in the middle Paranh River Valley. Verh. Int. Ver. Theor. Angew. Limnol. 17,

Boon, J . J., Wetzel, R. G., and Godshalk, G. L. (1982). Pyrolysis mass spectrometry of some

Boyd, C. E. (1970). Losses of mineral nutrients during decomposition of Typha latifolia.

Bray, J. R. (1963). Root production and the estimation of net productivity. Can. J. Bot. 41,

Brinson, M. M. (1977). Decomposition and nutrient exchange of litter in an alluvial swamp

155- 187.

359-362.

239-25 1.

munity. Ecology 53, 714-719.

Limnology.” Longman, London.

J. Bot. 55, 630-638.

81-96.

23, 261-263.

179-184.

1035-1050.

Scirpus species and their decomposition products. Limnol. Oceanogr. 27, 839-848.

Arch. Hydrobiol. 66, 51 1-517.

65-72.

forest. Ecology 58, 601-609.


Brinson, M. M., Lugo, A. E., and Brown, S. (1981). Primary productivity, decomposition and consumer activity in freshwater wetlands. Annu. Rev. Ecol. Syst. 12, 123-161.

Bristow, J. M. (1975). The structure and function of roots in aquatic vascular plants. In “The Development and Function of Roots” (J. G. Torrey and D. T. Clarkson, eds.), pp. 221- 236. Academic Press, New York.

Brown, D. S. (1961). The food of the larvae of Chloeon dipterum L. and Baetis rhodani (Pictet) (Insecta, Ephemeroptera). J. Anim. Ecol. 30, 55-75.

Bryant, R. J., Woods, L. E., Coleman, D. C., Fairbanks, B. C., McClellan, J. F., and Cole, C. V. (1982). Interactions of the bacterial and amoeba1 populations in soil microcosms with fluctuating moisture content. Appl. Environ. Microbiol. 43, 747-752.

Burgis, M. J., Darlington, J. P. E. C., Dunn, I. G., Ganf, G. G., Gwahaba, J. J., and McGowan, L. M. (1973). The biomass and distribution of organisms in Lake George, Uganda. Proc. R. SOC. London Ser. B. 184, 271-298.

Buttery, B. R., Williams, W. T., and Lambert, J. M. (1965). Competition between Glyceria maxima and Phragmites communis in the region of Surlingham Broad. J. Ecol. 53, 183- 195.

Calow, P. (1974). Evidence for bacterial feeding in Planorbis contortus Linn. (Gastropoda: Pulmonata). Proc. Malac. SOC. London 41, 145-156.

Calow, P., and Calow, L. J. (1975). Cellulase activity and niche separation in freshwater gastropods. Nature (London) 255, 478-480.

Carter, G. S. (1934). Results of the Cambridge Expedition to British Guiana, 1933. The freshwaters of the rain forest areas of British Guiana. J. Linn. SOC. London (Zool.) 39, 143- 193.

Carter, G. S., and Beadle, L. C. (1930). The fauna of the swamps of the Paraguayan Chaco in relation to its environment-I. Physicochemical nature of the environment. J. Linn. SOC. London (Zool.) 37, 205-258.

Chambers, M. R. (1971). The dominance, production and utilization of Gammarus tigrinus (Sexton) in the exposed Phragmites reed beds of the Tjeukemeer (Holland). Hidrobiologia

Chamie, J. P. and Richardson, C. J. (1978). Decomposition in northern wetlands. In “Fresh- water Wetlands” (R. E. Good, D. F. Whigham, and R. L. Simpson, eds.), pp. 115-130. Academic Press, New York.

Chesters, C. G. C. (1960). Certain problems associated with the decomposition of soil organic matter by fungi. In “The Ecology of Soil Fungi” (D. Parkinson and J. S. Waid, eds.), pp. 223-238. Liverpool University, Liverpool.

12, 297-303.

Conway, V. M. (1949). The bogs of central Minnesota. Ecol. Monogr. 19, 173-206. Crawford, D. L., and Crawford, R. L. (1976). Microbial degradation of lignocellulose: the

lignin component. Appl. Environ. Microbiol. 31, 714-717. Daget, J. (1952). Mtmoires sur la biologie des poissons du Niger Moyen. I-Biologie et crois-

sance des tsphces du genre Alestes. Bull. Inst. Fr. Afr. Noire Ser. A 14. Danell, K., and Andersson, A . (1982). Dry weight loss and colonization of plant litter by

macroinvertebrates: Plant species and lake types compared. Hydrobiologia 94,91-96. Danell, K., and Sjoberg, K. (1979). Decomposition of Carex and Equisetum in a northern

Swedish lake: Dry weight loss and colonisation by macro-invertebrates. J. Ecol. 67, 191- 200.

Davis, C. B., and van der Valk, A. G. (1978a). Litter decomposition in prairie glacial marshes. In “Freshwater Wetlands” (R. E. Good, D. F. Whigham, and R. L. Simpson, eds.), pp. 99-113. Academic Press, New York.

Davis, C. B., and van der Valk, A. G. (1978b). The decomposition of standing and fallen litter of Typha glauca and Scirpus fluviatilis. Can. J. Bot. 56, 662-675.


Davis, C. B., and van der Valk, A. G. (1984). Uptake and release of nutrients by living and

Davis, R. B. (1974). Tubificids alter profiles of redox potential and pH in a profundal lake

Day, F. P. (1982). Litter decomposition rates in the seasonally flooded Great Dismal Swamp.

de Haan, H. (1972). Some structural and ecological studies on soluble humic compounds from Tjeukemeer. Verh. Int. Ver. Theor. Angew. Limnol. 18, 685-695.

de Haan, H. (1974). Effect of a fulvic acid fraction on the growth of a Pseudomonas from Tjeukemeer. Freshwat. Biol. 4, 301-310.

de Jong, J., and Kok, T. (1978). The purification of wastewater and effluents using marsh vegetation and soils. In “5th International Symposium on Aquatic Weeds,” pp. 135-142. European Weed Research Society, Amsterdam.

den Hartog, C. (1978). Structural and functional aspects of macrophyte-dominated aquatic systems. In “5th International Symposium on Aquatic Weeds,” pp. 35-41. European Weed Research Society, Amsterdam.

Dobrowolski, K. A. (1973). Role of birds in Polish wetland ecosystems. Pol. Arch. Hy- drobiol. 20, 217-221.

Dokulil, M. (1973). Planktonic primary production within the Phragmites community of Lake Neusiedlersee (Austria). Pol. Arch. Hydrobiol. 20, 175-180.

Dokulil, M. (1975). Bacteria in the water and mud of Neusiedlersee (Austria). In “Limnology of Shallow Waters” (J. Saladki and J. E. Ponyi, eds.), pp. 135-140. Akadtmiai Kiad6, Budapest.

DvoHk, J. (1970a). Horizontal zonation of macrovegetation, water properties and macrofauna in a littoral stand of Glyceria aquatica (L.) Wahlb. in a pond in South Bohemia. Hydrobiologia 35, 17-30.

Dvofhk, J. (1970b). A quantitative study on the macrofauna of stands of emergent vegetation in a carp pond of South-West Bohemia. Rozpr. Cesk. Akad. Vpd, Rada Mat. PFir. Vpd.

Dykyjovh, D., and Hradeckh, D. (1973). Productivity of reed-bed stands in relation to the ecotype, microclimate and trophic conditions of the habitat. Pol. Arch. Hydrobiol. 20,

Edwards, R. W. (1958). The effect of Chironomus riparius Meigen on the redox potentials of settled activated sludge. Ann. Appl. Biol. 46, 457-464.

Egglishaw, H. J. (1964). The distributional relationships between the bottom fauna and plant detritus in streams. J. Anim. Ecol. 33, 463-470.

Egler, F. E. (1952). Southeast saline Everglades vegetation, Florida, and its management. Vegetatio 3, 213-265.

Farmer, V. C., and Morrison, R. I. (1964). Lignin in Sphagnum and Phragmites and in peats derived from these plants. Geochim. Cosmochim. Acta 28, 1537-1546.

Federle, J. W., and Vestal, J. R. (1980a). Microbial colonization and decomposition of Carex litter in an Arctic lake. Appl. Environ. Microbiol. 39, 888-893.

Federle, J. W., and Vestal, J. R. (1980b). Lignocellulose mineralization by Arctic lake sediments in response to nutrient manipulations. Appl. Environ. Microbiol. 40, 32-39.

Fenchel, T., and Harrison, P. G. (1976). The significance of bacterial grazing and mineral cycling for the decomposition of particulate detritus. In “The Role of Terrestrial and Aquatic Organisms in Decomposition Processes” (J. M. Anderson and A. Macfadyen, eds.), pp. 285-299. Blackwell, Oxford.

Fiala, K. (1973). Growth and production of underground organs of Typha angustifolia L., Typha latifolia L. and Phragmites communis Trin. Pol. Arch. Hydrobiol. 20, 59-66.

decomposing Typha glauca tissues at Eagle Lake, Iowa. Aquat. Bot. (in press).

sediment. Limnol. Oceanogr. 19, 342-346.

Ecology 63, 670-678.

80, 63-1 14.

111-119.


Fish, D., and Carpenter, S. R. (1982). Leaf litter and larval mosquito dynamics in tree-hole ecosystems. Ecology 63, 283-288.

Fish, G. R. (1956). Some aspects of the respiration of six species of fish from Uganda. J. Exp. Biol. 33, 186-195.

Fisher, S. G., and Carpenter, S. R. (1976). Ecosystem and macrophyte primary production of the Fort River, Massachusetts. Hydrobiologiu 47, 175-187.

Flenley, J. (1979). “The Equatorial Rain Forest: A Geological History.” Butterworth, Lon- don.

Furtado, J. I., and Verghese, S. (1981). Nutrient turnover in a freshwater inundated forested swamp, the Tasek Bera, Malaysia. Verh. Int. Ver. Theor. Angew. Lirnnol. 21, 1200- 1206.

Gajevskaya, N. S. (1969). “The Role of Higher Aquatic Plants in the Nutrition of the Animals of Fresh-water Basins” (Trans. by D. G. Maitland Muller, Ed. by K. H. Mann). 3 Volumes. National Lending Library for Science, Boston Spa, England.

Gaudet, J. (1976). Nutrient relationships in the detritus of a tropical swamp. Arch. Hydrobiol.

Gaudet, J. (1977). Uptake, accumulation, and loss of nutrients by papyrus in tropical swamps.

Gaudet, J. (1979). Seasonal changes in nutrients in a tropical swamp: North Swamp, Lake Naivasha, Kenya. J. Ecol. 67, 954-981.

Gerloff, G. C., and Krumbholz, P. H. (1966). Tissue analysis as a measure of nutrient availability for the growth of aquatic plants. Limnol. Oceunogr. 11, 529-537.

Given, P. A. (1975). Environmental organic chemistry of bogs, marshes and swamps. In “En- vironmental Chemistry,” Vol. 1. pp. 55-80. The Chemical Society, London.

Godlewska-Lipowa, W. A. (1975). Destruction of organic matter in the water of some Ma- surian lakes of varying trophism. In “Limnology of Shallow Waters” (J. Saladki and J. E. Ponyi, eds.), pp. 141-147. Akadtmiai Kiad6, Budapest.

Godshalk, G. L., and Wetzel, R. G. (1977). Decomposition of macrophytes and the metabolism of organic matter in sediments. In “Interactions between Sediments and Fresh Water” (H. L. Golterman, ed.), pp. 258-264. Junk, The Hague.

Godshalk, G. L.. and Wetzel, R. G. (1978a). Decomposition in the littoral zone of lakes. In “Freshwater Wetlands” (R. E. Good, D. F. Whigham, and R. L. Simpson. eds.), pp. 131- 143. Academic Press, New York.

Godshalk, G. L., and Wetzel, R. G. (1978b). Decomposition of aquatic angiosperms. I. Dis- solved components. 11. Particulate components. 111. Zosteru marina L. and a conceptual model of decomposition. Aquut. Bot. 5, 281-300, 301-327, 329-354.

Godwin, H. (1975). “The History of the British Flora.” 2nd ed. Cambridge University Press, London and New York.

Gorbunov, K. V. (1953). Raspad ostatkov vysshikh vodnykh rastenii i ego ekologicheskaia vol’v vodoemakh nizhnei zony Delty Volgi. Tr. Vses. Gidrobiol. Ova. 5, 158-202.

Gorham, E. (1953). Chemical studies on soils and vegetation of waterlogged habitats in the English Lake District. J. Ecol. 41, 345-360.

Gwahaba, J. J. (1975). The distribution, population density, and biomass of fish in an equatorial lake, Lake George, Uganda. Proc. R. SOC. London Ser. B. 190, 393-414.

Harding, D. (1960). Lake Kariba. the hydrology and development of the fisheries. In “Man- made Lakes” (R. H. Lowe-McConnell, ed.), pp. 7-18. Academic Press, New York.

Hargrave, B. T. (1970a). The effect of a deposit-feeding amphipod on the metabolism of benthic microflora. Limnol. Oceanogr. 15, 21-30.

Hargrave, B. T. (1970b). The utilization of benthic microflora by Hyulellu uztecu (Amphi- poda). J. Anim. Ecol. 39,427-437.

78, 213-239.

ECOIO~V 58,415-422.


Hargrave, B. T. (1972). Aerobic decomposition of sediment and detritus as a function of particle surface area and organic content. Limnol. Oceanogr. 17, 583-596.

Hargrave, B. T. (1975). The importance of total and mixed-layer depth in the supply of organic material to bottom communities. In “Limnology of Shallow Waters” (J. Salariki and J. E. Ponyi, eds.), pp. 157-165. Akadkmiai Kiadb, Budapest.

Hargrave, B. T. (1976). The central role of invertebrate faeces in sediment decomposition. In “The Role of Terrestrial and Aquatic Organisms in Decomposition Processes” (J. M. An- derson and A. Macfadyen, eds.), pp. 301-321. Blackwell, Oxford.

Harrison, E. S.. and Bloom, A. L. (1977). Sedimentation rates on tidal salt marshes in Con- necticut. J. Sediment. Petrol. 47, 1484-1490.

Harrison, M. J., Wright, R. J., and Morita, R. Y. (1971). Method for measuring mineralization in lake sediments. Appl. Microbiol. 21, 698-702.

Harrison. P. G., and Mann, K. H. (1975). Detritus formation from eelgrass (Zostera marina L.): The relative role of fragmentation, leaching and decay. Limnol. Oceanogr. 20, 924- 934.

Hayes, F. R., and Anthony, E. H. (1959). Lake water and sediment VI. Limnol. Oceanogr. 4, 299-3 15.

Herbst, G. N. (1982). Effect of leaf type on the consumption rate of aquatic detritivores.

Hickling, C. F. (1961). “Tropical Inland Fisheries.” Longman, London. Hill, B. H., and Webster, J. R. (1982). Aquatic macrophyte breakdown in an Appalachian

river. Hydrobiologia 89, 53-59. Ho, Y. B. (1979). Chemical composition studies on some aquatic macrophytes in three Scottish

lochs. I. Chlorophyll, ash, carbon, nitrogen and phosphorus. Hydrobiologia 63, 161-166. Ho, Y. B. (1980). Development of foliage structure in Phragmites australis (Cav.) Trin. ex

Steudel stands in Scottish lochs. Hydrobiologia 70, 159-164. Hobbie, J. E., Traaen, T., Rublee, P., Reed, J. P., McMiller, M. C., and Fenchel, T. (1980).

Decomposers, bacteria and microbenthos. In “Limnology of Tundra Ponds” (J. E. Hob- bie, ed.). USIIBP Synth. Ser. 13, 340-387.

Hodkinson, I. D. (1975a). Dry weight loss and chemical changes in vascular plant litter of terrestrial origin, occurring in a beaver pond ecosystem. J. Ecol. 63, 131-142.

Hodkinson, I. D. (1975b). A community analysis of the benthic insect fauna of an abandoned beaver pond. J. Anim. Ecol. 44, 533-551.

Hofsten, B. V., and Edberg, N. (1972). Estimating the rate of degradation of cellulose fibers in water. Oikos 23, 29-34.

Howard-Williams, C. (1972). Limnological studies in an African swamp: Seasonal and spatial changes in the swamps of Lake Chilwa, Malawi. Arch. Hydrobiol. 70, 379-391.

Howard-Williams, C. (1979a). The aquatic environment: 11. Chemical and physical characteristics of the Lake Chilwa swamps. In “Lake Chilwa” (M. Kalk, A. J. McLachlan, and C. Howard-Williams, eds.). Monogr. Biol. 35, 79-90.

Howard-Williams, C. (1979b). Interactions between swamp and lake. In “Lake Chilwa” (M. Kalk, A. J. McLachlan, and C. Howard-Williams, eds.). Monogr. Biol. 35, 231-245.

Howard-Williams, C., and Howard-Williams, W. (1978). Nutrient leaching from the swamp vegetation of Lake Chilwa, a shallow African lake. Aquat. Bot. 4, 257-267.

Howard-Williams, C., and Junk, W. J. (1977). The chemical composition of central Ama- zonian macrophytes with special reference to their role in the ecosystem. Arch. Hydrobiol. 79,446-464.

Howard-Williams, C., and Lenton, G. M. (1975). The role of the littoral zone in the functioning of a shallow tropical lake ecosystem. Freshwater Biol. 5, 445-459.

Hiirlimann, H. (1951). Zur Lebensgeschichte des Schilfs an den Ufern der Schweizer Seen. Beitr. Geobot. Landesaufn. Schweiz, Bern 30, 1-232.

Hydrobiologia 89, 77-87.


Hutchinson, G. E. (1975). “A Treatise on Limnology,” Vol. 3, Limnological Botany. Wiley.

Imhof, G. (1973). Aspects of energy flow by different food chains in a reed-bed. A review.

Ingold, C. T. (1943). Further observations on aquatic Hyphomycetes of decaying leaves. Trans.

Ingold, C. T. (1951). Aquatic Ascomycetes: Ceriospora caudaesut n.sp. and Ophiobolus ty-

Ingold, C. T., and Chapman, B. (1952). Aquatic Ascomycetes: Loramycusjuncicola Weston

Iversen, T. M. (1974). Ingestion and growth in Sericostoma personaturn (Trichoptera) in re-

Jsrgensen, E. 0. (1957). Diatom periodicity and silicon assimilation. Dan. Bot. Ark. 18,

Kajak, Z . (1972). Analysis of the influence of fish on benthos by the method of enclosures. In “Productivity Problems of Freshwaters” (Z. Kajak and A. Hillbricht-Ilkowska, eds.), pp. 781-793. Polish Scientific Publishers, Warsaw.

Kajak. Z.. and Dusoge, K. (1975a). Macrobenthos of Lake Taltowisko. Ekol. Pol. 23,295- 316.

Kajak, Z., and Dusoge, K. (1975b). Macrobenthos of Mikdajskie Lake. Ekol. Pol. 23, 437- 457.

Kajak, Z., Dusoge, K., and Stalicykowska, A. (1968). Influence of mutual relations of organisms, especially Chironomidae, in natural benthic communities, on their abundance. Ann. Zool. Fenn. 5 , 49-56.

Kamp-Nielsen, L. (1974). Mud-water exchange of phosphate and other ions in undisturbed sediment cores and factors affecting the exchange rates. Arch. Hydrobiol. 13, 218-237.

Kaushik, N. K., and Hynes, H. B. N. (1971). The fate of dead leaves that fall into streams. Arch. Hydrobiol. 68, 465-515.

Kelly, C. A., and Chynoweth, D. P. (1981). The contributions of temperature and of the input of organic matter in controlling rates of sediment methanogenesis. Limnol. Oceanogr. 26,

Kikuchi, T., and PCrbs, J. M. (1977). Consumer ecology of seagrass beds. In “Seagrass Eco- systems” (C. P. McRoy and C. Helfferich, eds.), pp. 147-193. Dekker, New York.

Kirby, J. M., Webster, J. R.. and Benfield, E. F. (1984). The role of shredders in detrital dynamics of permanent and temporary streams. In “Dynamics of Lotic Ecoystems” Proc. Symp. Savannah R. Ecol. Lab. 1980 (in press).

Klopatek, J. M., and Stearns, F. W. (1978). Primary productivity of emergent macrophytes in Wisconsin freshwater marsh ecosystem. Am. Midl. Nut. 100, 320-332.

Kolodziejczyk, A., and Martynuska, A. (1980). Lymnaea stagnalt (L.)-Feeding habits and production of faeces. Ekol. Pol. 28, 201-217.

KomPrkovP, J., and KomArek, J. (1975). Comparison of pelagial and littoral primary production in a South Bohemian fish pond (Czechoslovakia). In “Limnology of Shallow Waters” (J. Salariki and J. E. Ponyi, eds.), pp. 77-95. AkadCmiai Kiadd, Budapest.

Krashenninikova, S. A. (1958). Mikrobiologicheskie protsessy raslada vodnoi rastitel nosti v litorali Rybinskogo Vodokhranilishcha. Byull. Inst. Biol. Vodokhran., Akad. Nauk SSSR

Krzywosz, T., Krzywosz, W., and Radziej, J. (1980). The effect of grass carp, Ctenophar- yngodon idella (Val.), on aquatic vegetation and ichthyofauna of Lake Dgal Wielki. Ekol. Pol. 28, 433-450.

Kvet, J., and Hudec, K. (1971). Effects of grazing by Grey-lag Geese on reed swamp plant communities. Hidrobiologia 12, 351-359.

New York.

Pol. Arch. Hydrobiol. 20, 165-168.

Br. Mycol. SOC. 26, 104-114.

phae. Trans. Br. Mycol. SOC. 34, 210-215.

and L. macrospora n.sp. Trans. Br. Mycol. SOC. 35, 268-272.

lation to the nitrogen content of ingested leaves. Oikos 25, 278-282.

1-54.

891-897.

2, 3-6.


Ladle, M. (1974). Aquatic Crustacea. In “Biology of Plant Litter Decomposition” (C. H. Dickinson and G. J. F. Pugh, eds.), Vol. 2, pp. 593-608. Academic Press, New York.

Ladle, M., Bass, J. A. B., and Jenkin, W. E. (1972). Studies on production and food consumption by the larval Simuliidae (Diptera) of a chalk stream. Hydrobiologia 39,429-448.

Larsen, V. J. (1982). The effects of pre-drying and fragmentation on the leaching of nutrient elements and organic matter from Phragmites australis (Cav.) Trin. litter. Aquat. Bot. 14,

Larsen, V. J.. and Schierup, H.-H. (1981). Macrophyte cycling of zinc, copper, lead and cadmium in the littoral zone of a polluted and a non-polluted lake. 11. Seasonal changes in heavy metal content of above-ground biomass and decomposing leaves of Phragmites australis (Cav.) Trin. Aquat. Bot. 11, 211-230.

Lovley, D. R., and Hug, M. J. (1982). Intermediary metabolism of organic matter in the sediment of a eutrophic lake. Appl. Environ. Microbiol. 43, 552-560.

Lowe-McConnell, R. H. (1975). “Fish Communities in Tropical Fresh Waters.” Longman, London.

Lush, D. L., and Hynes, H. B. N. (1973). The formation of particles in freshwater leachates of dead leaves. Limnol. Oceanogr. 18, 968-977.

Macan, T. T., and Kitching. A. (1972). Some experiments with artificial substrates. Verh. Int. Ver. Theor. Angew. Limnol. 18, 213-220.

McConnell, W. J. (1968). Limnological effects of organic extracts of litter in a southwestern impoundment. Limnol. Oceanogr. 13, 343-349.

McKinley, V. C.. and Vestal, J. R. (1982). Effects of acid on plant litter decomposition in an Arctic lake. Appl. Environ. Microbiol. 43, 1188-1195.

McLachlan, A. J. (1974). Recovery of the mud substrate and its associated fauna following a dry phase in a tropical lake. Limnol. Oceanogr. 19, 74-83.

McLachlan, S. M. (1971). The rate of nutrient release from grass and dung following im- mersion in lake water. Hydrobiologia 37, 521-530.

Mann, K. H. (1956). A study of the oxygen consumption of five species of leech. J. Exp. Biol. 33, 615-626.

Marlier, G. (1958). Recherches hydrobiologiques du Lac Tumba. Hydrobiologia 10,352-385. Marlier, G . (1967). Ecological studies on some lakes of the Amazon Valley. Amazoniana 1,

Mason, C. F. (1976). Relative importance of fungi and bacteria in the decomposition of Phrag-

Mason, C. F. (1977). “Decomposition.” Arnold, London. Mason, C. F., and Bryant, R. J. (1974). The structure and diversity of the animal communities

in a broadland reed swamp. J. Zool. 172, 289-302. Mason, C. F., and Bryant, R. J. (1975a). Production, nutrient content and decomposition of

Phragmites communis Trin. and Typha angustifolio L. J. Ecol. 63, 11-95. Mason, C. F., and Bryant, R. J. (1975b). Periphyton production and grazing by chironomids

in Alderfen Broad, Norfolk. Freshwater Biol. 5, 271-277. Matthes, H. (1964). Les poissons du Lac Tumba et de la region d’Ikela. Etude systkmatique

et ecologique. Ann& Mus. r. AJr. Centr. (Sci. Zool.) 126. (204 pp.). Meschkat, A. (1934). Der Bewuchs in den Rohrichten des Plattensees. Arch. Hydrobiol. 27,

Minckley, W. L. (1963). The ecology of a spring stream Doe Run, Meade County, Kentucky.

Mizuno, T., and Mori, S. (1970). Preliminary hydrobiological survey of some Southeast Asian

Monakov, A. Y. (1972). Review of studies on feeding of aquatic invertebrates conducted at

29-39.

91-115.

mites leaves. Hydrobiologia 51, 65-69.

436-5 17.

Wildl. Monogr. 11 (124 pp.).

inland waters. Biol. J. Linn. SOC. 2 , 77-117.


the Institute of Biology of Inland Waters, Academy of Science, USSR. J. Fish. Res. Bd. Can. 29, 363-383.

Monk, D. C. (1976). The distribution of cellulase in freshwater invertebrates of different feeding habits. Freshwater Biol. 6, 471-475.

Monk, D. C. (1977). The digestion of cellulose and other dietary components, and the pH of the gut in the amphipod Gammaruspulex L . Freshwater Biol. 7 , 431-440.

Moore, P. D. (1980). Exploiting papyrus. Nature (London) 284, 510. Moore, P. D., and Bellamy, D. J. (1974). “Peatlands.” Elek Science, London. Morrison, S. J., and White, D. C. (1980). Effects of grazing by estuarine gammaridean am-

phipods on the microbiota of allochthonous detritus. Appl. Environ. Microbiol. 40, 659- 671.

Mortimer, C. H. (1941). The exchange of dissolved substances between mud and water in lakes. I and 11. J. Ecol. 29, 280-329; 30, 147-201.

Moss, B. (1970). Seston composition in two freshwater pools. Limnol. Oceanogr. 15, 504- 513.

Moss, B. (1979). The Lake Chilwa ecosystem-A limnological overview. In “Lake Chilwa” (M. Kalk, A. J. McLachlan, and C. Howard-Williams, eds.). Monogr. Biol. 35, 399-415.

Moss, B. (1980). The relative importance of different producing communities. In “The Func- tioning of Freshwater Ecosystems” (E. D. LeCren and R. H. Lowe-McConnell, eds.), pp. 214-215. Cambridge University Press, London and New York.

Moss, B., and Moss, J. (1969). Aspects of the limnology of an endorheic African lake (L. Chilwa, Malawi). Ecology 50, 109-1 18.

Odum, W. E., and Heywood, M. A. (1978). Decomposition of intertidal freshwater marsh plants. In “Freshwater Wetlands” (R. E. Good, D. F. Whigham, and R. L. Simpson, eds.), pp. 89-97. Academic Press, New York.

Olfih, J. (1969a). The quantity, vertical and horizontal distribution of the total bacterioplankton of Lake Balaton in 1966/7. Ann. Biol. (Tihany) 36, 185-195.

OMh, J. (1969b). A quantitative study of the saprophytic and total bacterioplankton in free open water and free littoral zone of Lake Balaton in 1968. Ann. Biol. (Tihany) 36, 197- 212.

OIBh, J. (1972). Leaching, colonization and stabilization during detritus formation. Mem. 1st. ital. Idrobiol. (Suppl.) 29, 105- 127.

Onuf, C. P., Teal, J. M., and Valiela, I. (1977). Interactions of nutrients, plant growth and herbivory in a mangrove ecosystem. Ecology 58, 514-526.

Opalifiski, K. W. (1971). Macrofauna communities of the littoral of Mikdajskie Lake. Pol. Arch. Hydrobiol. 18, 275-285.

Otsuki, A., and Wetzel, R. G. (1973). Interactions of yellow organic acids with calcium carbonate in freshwater. Limnol. Oceanogr. 18, 490-493.

Otsuki, A., and Wetzel, R. G. (1974). Release of dissolved organic matter by autolysis of a submerged macrophyte, Scirpus subterminalis. Limnol. Oceanogr. 19, 842-845.

Paerl, H. W. (1977). Bacterial sediment formation in lakes: Trophic implications. In “Inter- actions between Sediments and Fresh Water” (H. L. Golterman, ed.), pp. 40-47. Junk, The Hague.

Paul, R. W., Benfield, E. F., and Cairns, J. (1978). Effects of thermal discharges on leaf decomposition in a river ecosystem. Verh. Int. Ver. Theor. Angew. Limnol. 20, 1759- 1766.

Pelikfin, J., Svoboda, J., and KvM, J. (1971). Relationships between the population of musk- rats (Ondutra zibethica) and the primary production of cattail (Typha latifolia). Hidro- biologia 12, 177-180.

Pelikfin, J., Hudec, K., and SlastnJr, K. (1978). Animal populations in fishpond littorals. In


“Pond Littoral Ecosystems” (D. Dykyjovii and J. Kvet, eds.), pp. 74-79. Springer-Verlag, Berlin and New York.

Peterson, B. J., Hobbie, J. E., and Haney, J. F. (1978). Duphniu grazing on natural bacteria. Limnol. Oceunogr. 23, 1039-1044.

Pieczydska, E. (1972). Production and decomposition in the littoral zone of lakes. In “Pro- ductivity Problems of Freshwaters” (Z. Kajak and A. Hillbricht-Ilkowska, eds.), pp. 271- 285. Polish Scientific Publishers, Warsaw.

Pieczyhska, E. (1973). Experimentally increased fish stock in the pond type Lake Warniak. XI. Food resources and availability of the eulittoral zone for fish. Ekol. Pol. 21, 583-593.

Pieczyriska, E. (1975). Ecological interactions between land and the littoral zones of lakes (Masurian Lakeland, Poland). In “Coupling of Land and Water Systems” (A. D. Hasler, ed.), pp. 263-276. Springer-Verlag, Berlin and New York.

Pieczydska, E. (1977). Numbers and biomass of the littoral fauna in Mikdajskie Lake and in other Masurian lakes. Ekol. Pol. 25, 45-57.

Pieczyhska, E., and Ozimek, T. (1976). Ecological significance of lake macrophytes. Int. J. Ecol. Environ. Sci. 2, 115-128.

Trin. Pol. Arch. Hydrobiol. 17, 357-362. Planter, M. (1973). Physical and chemical conditions in the helophytes zone of the lake littoral.

Pol. Arch. Hydrobiol. 20, 1-7. Polisini, J. M., and Boyd, C. E. (1972). Relationships between cell-wall fractions, nitrogen,

and standing crop in aquatic macrophytes. Ecology 53, 484-488. Polunin, N. V. C. (1982a). Seasonal variation in seston and organic matter accumulation in

a sheltered fenland pond. Hydrobiologiu 94, 155-162. Polunin, N. V. C. (1982b). Processes contributing to the decay of reed (Phrugmites uustrulis)

litter in fresh water. Arch. Hydrobiol. 94, 182-209. Polunin, N. V. C. (1982~). Effects of the freshwater gastropod Plunorbis curinutus on reed

(Phrugmites uustrulis) Litter microbial activity in an experimental system. Freshwater Biol.

Prus, T. (1976). Experimental and field studies on ecological energetics of Asellus aquaticus L. (Isopoda) I. Assimilability of lipids, proteins and carbohydrates. Ekol. Pol. 24, 461- 472.

Prus. T. (1977). Experimental and field studies on ecological energetics of Asellus aquaticus L. (Isopoda) IV. Energy budget of a population in the littoral zone of Powsidkie Lake.

Pugh, G. J. F., and Mulder, J. L. (1971). Mycoflora associated with Typhu latifolio. Trans.

Reddy, K. R., and Patrick, W. H. (1975). Effect of alternate aerobic and anaerobic conditions on redox potential, organic matter decomposition and nitrogen loss in flooded soil. Soil Biol. Biochem. 7, 87-94.

Reed, F. C. (1979). Decomposition of Acer rubrum leaves at three depths in an eutrophic Ohio lake. Hydrobiologiu 64, 195-197.

Rich, P. H., and Wetzel. R. 0. (1978). Detritus in the lake ecosystem. Am. Nut. 112, 57- 71.

Richards, F. J. (1934). The salt marshes of the Dovey Estuary: IV. The rates of vertical accretion, horizontal extension and scarp erosion. Ann. Bot. 48, 225-259.

Robertson, M. L., Mills, A. L., and Zieman. J. C. (1982). Microbial synthesis of detritus- like particles from dissolved organic carbon released by tropical seagrasses. Mur. Ecol. Prog. Ser. I , 279-285.

Planter, M. (1970). Elution of mineral components our of dead reed Phrugmiia communb

12, 547-552.

EkOl. POI. 25, 593-623.

Br. MYCO~. SOC. 57, 273-282.


Rodewald-Rudescu, L. (1974). “Das Schilfrohr.” (Die Binnengewasser, 27). Schweizerbart’

Roos, P. J. (1982). Fate and final distribution of dead reed stems in Lake Marsseveen. Hy-

Rudescu, L., Niculescu, C., and Chivu, I. P. (1965). Monogr. Stufului din Delta Dunarii,

Russell-Hunter, W. D. (1970). “Aquatic Productivity.” Macmillan, New York. Rybak, J. I. (1969). Bottom sediments of the lakes of various trophic type. Ekol. Pol. 17,

Rzdska, J. (1974). The Upper Nile swamps, a tropical wetland study. Freshwater Biol. 4,

Sandon, H., and el Thayib, A. (1953). The food of some common Nile fish. Sudan Notes Rec. 34, 205-229.

Saunders, G. W. (1972). The transformation of artificial detritus in lake water. Mem. Ist. Ital. Idrobiol. (Suppl.) 29, 261-288.

Saunders, G. W. (1976). Decomposition in freshwater. In “The Role of Terrestrial and Aquatic Organisms in Decomposition Processes” (J. M. Anderson and A. Macfadyen, eds.), pp. 341-373. Blackwell, Oxford.

Saunders, G. W., Cummins, K. W., Gak, D. Z., Pieczyliska, E., StraSkrabovB, V., and Wetzel, R. G. (1980). Organic matter and decomposers. In “The Functioning of Freshwater Eco- systems” (E. D. LeCren and R. H. Lowe-McConnell, eds.), Int. Biol. Programme 22,341- 392.

Schroder, R. (1975). Release of plant nutrients from reed borders and their transport into the open waters of the Bodensee-Untersee. Symp. Biol. Hung. 15, 21-27.

Sculthorpe, C. D. (1967). “The Biology of Aquatic Vascular Plants.” Arnold, London. Shaver, G. R., Chapin, F. S., and Billings, W. D. (1979). Ecotypic differentiation in Carex

aquatilis on ice-wedge polygons in the Alaskan coastal tundra. J. Ecol. 67, 1025-1046. Sherr, B. F., Sherr, E. B., and Berman, T. (1982). Decomposition of organic detritus: A

selective role for microflagellate Protozoa. Limnol. Oceanogr. 25, 765-769. Sibert, J., Brown, T. J., Healey, M. C.. Kask, B. A., and Naiman, R. J. (1977). Detritus-

based food webs: Exploitation by juvenile chum salmon (Oncorhynchus keta). Science 196,

Sircar, S. S. G., De, S. C., and Bhowmick, H. D. (1940). Microbiological decomposition of plant materials. I. Changes in the constituents of rice straw (Kanaktara) produced by microorganisms present in soil suspension under aerobic, anaerobic and water-logged conditions. Indian J. Agric. Sci. 10, 119-151.

Sirjola, E. (1969). Aquatic vegetation of the River Teuronjoki, South Finland, and its relation to water velocity, Ann. Bot. Fenn. 6, 68-75.

Skuhravl, V. (1978). Invertebrates: Destroyers of common reed. In “Pond Littoral Ecosys- tems” (D. DykyjovB and J. K v l , eds.), pp. 376-388. Springer-Verlag, Berlin and New York.

sche, Stuttgart.

drobiol. Bull. 16, 113-1 14.

Editura Academiei Republicii Socialiste Romania, Bucharest.

61 1-662.

1-30.

649-650.

Smirnov, N. N. (1961). Food cycles in Sphagnum bogs. Hydrobiologia 17, 175-182. Smirnov, N. N. (1962). On nutrition of caddis worms Phryganea grandis L. Hydrobiologia

Smork, L. A.. and Stoneburner, D. L. (1980). The response of aquatic macroinvertebrates to

Smyly, W. J. P. (1955). Comparison of the Entomostraca of two artificial moorland ponds

Sorensen, L. H. (1974). Rate of decomposition of organic matter in soil as influenced by

19, 252-261.

aquatic macrophyte decomposition. Oikos 35, 397-403.

near Windermere. Verh. Int. Ver. Theor. Angew. Limnol. 12, 421-424.


repeated air drying-wetting and repeated additions of organic material. Soil Biol. Biochem.

Soszka, G. J. (1975). Ecological relations between invertebrates and submerged macrophytes

Spence, D. H. N. (1967). Factors controlling the distribution of freshwater macrophytes with

Spence, D. H. N. (1982). The zonation of plants in freshwater lakes. Adv. Ecol. Res. 12, 37-

Stevenson, I. J. (1956). Some observations on the microbial activity of remoistened air-dried

Street, M., and Titmus, G. (1982). A field experiment on the value of allochthonous straw as

Stubbs, J. M. (1949). Freshwater fisheries in the northern Bahr el Ghazal. Sudan Notes Rec.

Suberkropp, K., Godshalk, G. L., and Klug, M. J. (1976a). Fungi and bacteria associated

Suberkropp, K., Godshalk, G. L., and Klug, M. J. (1976b). Changes in the chemical com-

Swift, M. J., Heal, 0. W., and Anderson, J. M. (1979). “Decomposition in Terrestrial Eco-

Szajnowski, F. (1970). The relations between the reed standing-crop and fishery effect. Pol.

Szczepadska, W., and Szczepadski, A. (1973). Emergent macrophytes and their role in wetland

Szczepadski, A. (1978). Ecology of macrophytes in wetlands. Pol. Ecol. Stud. 4, 45-94. Taligoola, T. K., Apinis, A. E., and Chesters, C. G. C. (1972). Microfungi colonizing col-

lapsed aerial parts of Phragmites communis Trin. in water. Nova Hedwigia 23, 465-472. Talling, J. F. (1957). The longitudinal succession of the water characteristics in the White

Nile. Hydrobiologia 11, 73-89. Tansley, A. G. (1939). “The British Islands and their Vegetation.” Cambridge University

Press, London and New York. Thomas, I. F. (1961). The Cladocera of the swamps of Uganda. Crustaceana 2, 108-125. Thompson, K. (1977). The primary productivity of African wetlands, with particular reference

to the Okavango Delta. In “Proceedings of the Symposium on the Okavango Delta and its Future Utilization,” pp. 67-69. Botswana Society, Gaborone.

Tobler, F. (1943). Stengelbau, Festigkeits- und Verwertungsunterschiede beim Schilfrohr (Phragmites communis Trin.). Angew. Bot. 25, 165-177.

Triska, F. J., and Sedell, J. R. (1976). Decomposition of four species of leaf litter in response to nitrate manipulation. Ecology 57, 783-792.

Ulehlovd, B (1978a). Decomposers in the fish pond littoral ecosystem. In “Pond Littoral Eco- systems” (D. Dykyjovd and J. KvEt, eds.), pp. 80-87. Springer-Verlag, Berlin and New York.

Ulehlovd, B. (1978b). Decomposition processes in the fish pond littoral. In “Pond Littoral Ecosystems” (D. DykyjovB and J. Kv&, eds.), pp. 341-353. Springer-Verlag, Berlin and New York.

Ulehlovd, B., HusBk, S., and Dvofdk, J. (1973). Mineral cycles in reed stands of Nesyt fishpond in southern Moravia. Pol. Arch. Hydrobiol. 20, 121-129.

Uotila, P. (1971). Distribution and ecological features of hydrophytes in the polluted Lake Vanajavesi, S . Finland. Ann. Bot. Fenn. 8, 257-295.

6, 287-292.

in the lake littoral. Ekol. Pol. 23, 393-415.

particular reference to the lochs of Scotland. J. Ecol. 55, 147-170.

125.

soils. Plant Soil 8 , 170-182.

food and substratum for lake macro-invertebrates. Freshwater Biol. 12, 403-410.

30, 245-25 1.

with leaves during processing in a woodland stream. Ecology 57, 707-719.

position of leaves during processing in a woodland stream. Ecology 57, 720-727.

systems.” Blackwell, Oxford.

Arch. Hydrobiol. 17, 363-371.

ecosystems. Pol. Arch. Hydrobiol. 20, 41-50.


Vaatanen, P., and Sundquist, J. (1977). Microbial cellulolytic activity of the brackish water of the Tviirminne Area, North Baltic Sea. Int. Revue Gesamten Hydrobiol. 62, 797-804.

Vallentyne. J. R. (1957). The molecular nature of organic matter in lakes and oceans, with lesser reference to sewage and terrestrial soils. J. Fish. Res. Bd. Can. 14, 33-82.

van Schreven, D. A. (1967). The effect of intermittent drying and wetting of a calcareous soil on carbon and nitrogen mineralization. Plant Soil 26, 14-32.

Visser, S. A. (1964). A study in the decomposition of Cypenrs papyrus in the swamps of Uganda, in natural peat deposits, as well as in the presence of various additives. East 4fr. Agric. For. J. 29, 268-287.

Voigts, D. K. (1976). Aquatic invertebrate abundance in relation to changing marsh vegetation. Am. Midl. Nut. 95, 313-322.

Voshell, J. R., and Simmons, G. M. (1977). An evaluation of artificial substrates for sampling macrobenthos in reservoirs. Hydrobiologia 53, 257-269.

Waksman, S. A., and Tenney, F. G. (1928). Composition of natural organic materials and their decomposition in the soil. 111. The influence of nature of plant upon the rapidity of its decomposition. Soil Sci. 26, 155-171.

Walker, D. (1970). Direction and rate in the British post-glacial hydroseres. In “Studies in the Vegetational History of the British Isles” (D. Walker and R. G. West, eds.), pp. 117- 139. Cambridge University Press, London and New York.

Wallace, J. B., and Merritt. R. W. (1980). Filtering ecology of aquatic insects. Annu. Rev. Entomol. 25, 103-132.

Wallace, J. B., Webster, J. R., and Cuffney, T. F. (1982). Stream detritus dynamics: Regu- lation by invertebrate consumers. Oecologia 53, 197-200.

Ward, G. M., and Cummins, K. W. (1979). Effects of food quality on growth of a stream detritivore, Paratendipes albimanus (Meigen) (Diptera, Chironomidae). Ecology 60,57-64.

Webster, J. R., and Simmons, G. M. (1978). Leaf breakdown and invertebrate colonization on a reservoir bottom. Verh. Int. Ver. Theor. Angew. Limnol. 20, 1587-1596.

Welcomme, R. L. (1970). Studies on the effects of abnormally high water levels on the ecology of fish in certain shallow regions of Lake Victoria. J. Zool. 160, 405-436.

Welcomme, R. L. (1979). “Fisheries Ecology of Floodplain Rivers.” Longman, London. Westlake, D. F. (1973). Aquatic macrophytes in rivers. A review. Pol. Arch. Hydrobiol. 20,

31-40. Wetzel, R. G. (1969). Factors influencing photosynthesis and excretion of dissolved matter by

aquatic macrophytes in hard-water lakes. Verh. Int. Ver. Theor. Angew. Limnol. 17,

-

72-85. Wetzel, R. G. (1975). “Limnology.” Sanders. Philadelphia. Wetzel, R. G., and Hough, R. A. (1973). Productivity and role of aquatic macrophytes in

lakes. An assessment. Pol. Arch. Hydrobiol. 20, 9-19. Wetzel, R. G., and Manny, B. A. (1972). Decomposition of dissolved organic carbon and

nitrogen compounds from leaves in an experimental hard-water stream. Limnol. Oceanogr.

Wetzel. R. G. , and Otsuki, A. (1973). Allochthonous organic carbon of a marl lake. Arch. Hydrobiol. 73, 31-56.

Wetzel, R. G., Rich, P. H., Miller. M. C., and Allen, H. L. (1972). Metabolism of dissolved and particulate detrital carbon in a temperate hard-water lake. Mem. Ist. Ital. Idrobiol.

Wiebe, W. J. (1979). Anaerobic benthic processes: Changes from the estuary into the conti- nental shelf. I n “Ecological Processes in Coastal and Marine Systems” (R. J. Livingston, ed.), pp. 469-485. Plenum, New York.

17, 927-93 1.

(SUPPI.) 29, 185-243.


Wiebe, W. J., and Pomeroy, L. R. (1972). Microorganisms and their association with aggre- gates and detritus in the sea: A microscopic study. Mem. Ist. Ital. Idrobiol. (Suppl.) 29,

Williams, R. (1972). Relationships between the water levels and the fish catches in Lakes Mweru and Mweru Na Ntipa, Zambia. 4fr . J. Trop. Hydrobiol. Fbh. 1(2), 21-32.

Winterbourn, M. J. (1978). An evaluation of the mesh bag method for studying leaf colonization by stream invertebrates. Verh. Int. Ver. Theor. Angew. Limnol. 20, 1557-1561.

Winterbourn, M. J., and Davis, S. F. (1976). Ecological role of Zelundopsyche ingens (Tri- choptera: Oeconesidae) in a beech forest stream ecosystem. Aust. J. Mar. Freshwater Res.

Zdanowski, M. K. (1977). Microbial degradation of cellulose under natural conditions. A

325-352.

27, 197-215.

review. Pol. Arch. Hydrobiol. 24, 215-225.

[advances in ecological research] advances in ecological research volume 14 volume 14 || the...

Documents