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SC. (I 990) /. Ecol. 78, 457-469 19 Weiner, 1. and Thomas, S.C. (1992) Ecology 739648-656 20 lurik, T.W. (1991) Oecologia 87, 539-550 21 BaliarC, CL., SBnchez, R.A., Scope), A.L.. Casal, I.). and Chersa. C.M. (1987) P/ant Cell Environ. IO, 551-557 22 Ballare, C.L.. Scopel, A.L. and SBnchez, R.A. (1990) Science 247, 329-332 23 Smith, H., Casal, 1.1. and lackson, G.M. (1991) Plant Cell Environ. 13, 73-78 24 Balla& C.L., Scopel, A.L., Slnchez, R.A. and Ghersa, C.M. (1988) Oecologia 76288-293 25 Ballare, CL.. Scopel, A.L. and Sgnchez, R.A. (1991) PlantCell Environ 14, 57-65

26 Ballark, CL., Scopel, A.L. and Sgnchez, R.A. (1989) P/ant Pbysiol. 89, 1324-l 330 27 Ballare. CL., Scopel, A.L. and Sgnchez, R.A. (1991) Oecologia 86, 561-567 28 Deregibus, V.A.. SPnchez, R.A., Casal, 1.1.

and Trlica, 1.1. (1985) J. App/. Eco/. 22 I, 199-206 29 Novoplansky, A., Cohen, D. and Sachs, T. ( 1990) Oecologia 82,490-493 30 Casal, ].I. and Smith, H. (1989) P/ant Ce// Environ. 12, 855-862 31 Weller, D.E. (1987) Eco/ogy68,813-821 32 Raynal, D.I. and Bazzaz, F.A. (1975) Ecology 56,35-49 33 Rees, M. and Long, M.I. (1992) Am. Nat. I39,484-508 34 Thompson, I.D. (1991) Trends Ecol. Evol. 6, 246-249 35 Weiner, J. and Thomas, SC. (1986) Oikos 47, 21 l-222 36 Weiner, 1. (1990) Trends Eco/. Eva/. 5, 360-364 37 Schmitt, 1. and Ehrhardt, D.W. (1990) Evolution 44,269-278 38 Wulff, R.D. (1986) 1. Ecol. 74, 115-121

39 Givnish, T.). (1982) Am. Nat. 120, 353-381 40 Geber, M.A. (1990) Evolution 44, 799-818 41 Lande, R. and Arnold, S.I. t 1983) Evolution37, l2lC-1226 42 Mitchell-Olds, T. and Shaw, R.G. (1987) Evolution 41, 1149-l I61 43 Wade, M.1. and Kalisz, S. (1990) Evolution 44, 1947-1955 44 Via, S. and Lande, R. (1985) Evolution 39. 505-522 45 Adamse, P., Kendrick, R.E. and Koornneef, M. (1988) Photochem. Photobiol. 48,833-84 I 46 Casal, I.]. (1988) Ann. App/. &Go/. I 12, 167-173 47 Solangaarachchi, S.M. and Harper, I.L. (1987) Oecologia 72, 372-376 48 McGraw, J.B. and Wulff. R.D. (1983) J. Theor. Biol. 103, 21-28 49 Wulff, R. (1989) Am. /. Bof 76, 131-132

Disturbance of the Phosphorus Cycle: A Case of Indirect

Effects of Human Activity

tiosphorus (P) often limits primary pro- ductivity of aquatic systems. Humans Grave altered the P cycle in aquatic sys- tems, directly, by mining P-rich rocli, and indirectly, through tke manipulation of other element cycles and tke alteration of aquatic food webs. Aquatic ecologists are becoming increasingly awure of tGre import- ance of these indirect alterations to bio- geochemical cycles. Quantitative predic- tions of these indirect effects will be dn important focus of future studies.

Humans have caused severe alterations to biogeochemical cyclesi,2. While some of these alter- ations are relatively direct, others are very indirect, being mediated through several interacting biologi- cal or chemical changes. However, the indirect effects of humans on biogeochemical cycles can be as severe as the direct effects and owing to the complex interactions involved, they are often overlooked. The phosphorus cycle exemplifies a biogeochemical cycle that has been severely altered by human activity.

Phosphorus cycling and direct human impact

Phosphorus is a critical part of liv- ing matter. It is contained in phos- pholipids of cell membranes, in ATP and in DNA. The high energy bond in PO, is crucial to the maintenance of the animated world3. Whether P

Nina Caraco is at the institute of Ecosystem Studies, The Mary Flagler Cary Arboretum, Box AB, Millbrook, NY 12545, USA.

0 1993, Elsewer Science Publishers Ltd (UK)

Nina F. Caraco

defines life itself or not, all organ- isms require P and autotrophic organisms must obtain this P from the environment. Acquisition of P can be energetically expensive because of the extremely low con- centration of available P found in many environments. For example, in a typical freshwater lake the avail- able P concentrations in the sur- rounding water is 0.0001 times that in the typical cell. For comparison, carbon, sulfur and nitrogen are on average 2.5 to 100 times more con- centrated in these same freshwater environments with respect to cell requirements4.

The importance of P to cell func- tions, coupled with its relatively low availability, is the reason why P fertilization can enhance auto- trophic production. Thus, P is a criti- cal component of plant fertilizers. The mining of P-rich rock to pro- duce these fertilizers has provided a large quantity of available P for agricultural land. Much of this P eventually reaches lakes, streams and ultimately the sea via rivers (Fig. 1). It has been estimated that on a global scale the mining of P- rich rock has more than doubled the supply of P to coastal seas, as

compared to preindustrial levelsi. In heavily populated and agricul- tural areas this effect is greater and many lakes have P-loading 10-l 00 times greater than preindustrial levels. This enhanced P-loading has led to increased productivity in many aquatic systems, resulting in some becoming eutrophic or hypereutrophic. Hypereutrophic waters are often characterized by foul smelling algal scums, hypoxic conditions (which lead to fish death) and blooms of poisonous algae. Thus, the direct effect of human activity on the P cycle is severe, and in many cases obvious and unap- pealing. Therefore, these direct effects have been the subject of much research and policy attention6.

Indirect human impact Phosphorus is an extremely

active element biologically and chemically. Chemically it interacts strongly with the iron, aluminium and calcium cycles. Biologically it is taken up directly from the environ- ment by not only photoautotrophs but also by chemotrophs and het- erotrophic bacteria and fungi. Given these multiple interactions it is not surprising that human activity

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I Phosphorus-containing rock I

Weathering

Undeveloped environment

1

Fertilizer Detergent ,

Agricultural Urban environment environment

Phosphorus in water bodies

Fig. I. Direct effects of humans on the phosphorus (P) cycle in aquatic systems. Mined P is converted to fertil- izers and made into detergent additives. This fertilizer P is largely used in agricultural landscapes where it can either accumulate in the soil and leach into waterways or be transported to cities as food. In the urban environ- ment large amounts of P are imported in the form of detergent and food. This P eventually reaches waterways primarily as sewage. The above redistribution of P from rock sources to agricultural and urban landscapes causes far greater run off from these areas than from undevel- oped landscapes4.

has and will continue to lead to indirect alterations of the P cycle. Recently these alterations have become more fully appreciated, both by ecologists and managers.

Below I review some examples that demonstrate the complex inter- actions that can occur between a human alteration and an ecosystem effect. These examples also demon- strate two fairly recent trends in the field of aquatic biogeochemistry: first, they emphasize the growing awareness of the importance of linked element interactions7; and secondly, they show the increasing awareness of the importance of macroorganisms in the regulation of element cycle$. Traditionally these element cycles have been portrayed as controlled by microbial reactions9.

Nitrate additions to aquatic systems

Nitrate inputs to lakes, estuaries and coastal waters have been increasing dramatically due to human activities. The increase in nitrate input comes from rain, groundwater and surface run off”. However, the source of much of this nitrate is fertilizer and internal combustion engines5. In coastal

waters this nitrate is a direct con- cern, as primary production in these waters is often nitrogen- limited and increased nitrogen (N) loading can result in eutro- phication”. In freshwater lakes the secondary effects of increased N input have been more of a concern.

In lakes, the P cycle is closely linked to the cycling of iron and oxygen (Fig. 2); nitrate can affect P cycling in lakes through its influ- ence on iron cycling. Nitrate can inhibit the reduction of oxidized iron (Fe3+) to the reduced form (Fe*+). Thus, if nitrate is present, Fe7+ levels can be maintained even in the absence of oxygen. As Fe” (but not Fe*+) is important in pre- venting P release from sediments to overlying waters, the presence of nitrate can indirectly inhibit P release from sediments’*. Evidence for the importance of this link comes from observed empirical relations between nitrate concen- trations in lakes and P loading from sediments of lakes13. Additionally, this relationship has been used to manage P release from sediments: in the lake management practice of sediment oxidation, nitrate is injected into surface sediments to prevent P release6.

The inhibitory effect of nitrate on P release from sediments relies on maintaining nitrate levels in the lake throughout the summer. If ni- trate levels are seasonally depleted (due to autotrophic uptake or dissimilatory reduction), the effect of nitrate is no longer to inhibit P release. In fact, it has been shown that the previously high nitrate con- centrations may actually enhance P release14. The mechanism for this phenomenon is not yet certain. One hypothesis is that the same bacteria that reduce nitrate (deni- trifying bacteria) will reduce Fe3+ in the absence of nitrate. High nitrate loading favors high populations of these bacteria which deplete the nitrate then switch to Fe’+ as an electron acceptor. The reduction of Fe3+, then, releases associated sorbed P. Thus, the effect of nitrate concentration on P release from sediments may be seasonally dependent15. In many lakes, increased nitrate loading may result in reduced P release from sedi- ments in the spring and early sum- mer but enhance P release in late

summer. This seasonal shift in P release may in turn be affecting seasonal succession of phytoplank- ton species and, thus, altering the entire planktonic community.

Sulfate loading to lakes Humans have altered the sulfur

(S) cycle dramatically by burning of fossil fuels (mainly coal) and using high-S fertilizer. Much of the S from these sources eventually reaches surface waters. The increased S loading to freshwater lakes has been suggested to directly affect primary production in a few low-S systems’6. However, the indirect effects of enhanced S (as sulfate) loading to lakes may be more widespread.

Like nitrate, sulfate may influ- ence P cycling in lakes through its interaction with the lacustrine iron cycle. Sulfate may affect both the pool of Fe3+ and the sorption of P to this Fe3+ pool (Fig. 2). The pool of Fe*+ depends not only on the dissolution of this pool but also on its formation. Sulfide, formed from the microbial reduction of sulfate, can decrease the rate of Fe” pro- duction by shunting Fe*+ from the ‘ferrous whee1’17 (Fig. 2). In addition to decreasing the pool of Fe”, sul- fate may decrease the sorption of phosphate-P to the remaining pool of Fe3+ due to anion competition for sorption sites. Due to both these effects, sulfate may decrease the efficiency of P trapping by Fe3+ in aquatic systemsIB. This decreased trapping allows more P to leak to surface waters where it can be used by phytoplankton. Thus, sulfate loading to lakes may be affecting productivity of lakes through its indirect effects on the P cycle.

H+ input to lakes The burning of fossil fuels and

use of internal combustion engines has increased the H’ content of rain (decrease in the pH) compared to that in pristine areas19. Sensitive regions are those that, due to their geology, have little ability to buffer H+ inputs, the result being the well known phenomenon of acidification of soils, groundwater, streams and lakes. One alteration that has been observed in some lakes in sensitive areas is an apparent ‘oligo- trophication’ or decrease in pro- ductivity20. There are many theories on how this oligotrophication might

52

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occur, and some of these concern the effects that H+ can have on P cycling in aquatic systems.

Increased H+ content of lakes in sensitive regions may be acting to enhance the binding of P by sedi- ments and decrease the supply of P to the water column. The mech- anism of this increased binding may be related in part to the fact that the chemical sorption of P on to Fe’+ or aluminium oxide surfaces is highly pH dependent (sorption is maximum under acid conditions, near pH of 4.5)2’,L2. Therefore, if the H+ content of a lake is increased, the potential exists for increased binding by sediments. With this increased binding there is less P leaked to surface waters where it can be used by phytoplankton. Thus, H’ increases in lakes can indi- rectly cause a reduction in produc- tivity of systems due to the indirect interaction with the P cycle.

fisCr introduction/rernovaI The introduction or removal of

species can have a dramatic effect on aquatic systems. Although some of these changes are due to the high abundance that introduced species can reach, the introduction of predatory species can often result in significant changes despite relatively low numbers and bio- mass. In lakes, the effects of intro- duced piscivorous fish have been well studied in recent years. The paradigm by which the introduction of predatory fish can influence lower trophic levels (eventually autotrophic organisms)23 has come to be known as the ‘trophic cascade’24.

Observations in lakes suggest that changes in fish biomass can lead to changes in the abundance of phytoplankton23-25. The mech- anism behind this link is often viewed solely as a response to trophic and competitive interac- tions. However, the possible alter- ation of biogeochemical cycles at the ecosystem level is becoming apparent. There are several ways in which food-web changes can alter P biogeochemistry (Box 1). The alter- ation of the P cycle can occur both by redistribution of P reservoirs within photic surface waters and by changes of movement of P into or out of surface waters. These alter- ations show that macroorganisms that have little direct impact on

energy flow in aquatic systems can still affect biogeochemical cycles. The widespread appreciation of the importance of macroorganisms in controlling biogeochemical cycles represents a relatively new direction in aquatic ecology.

Recent progress and future directions Quantifying indirect effects

Phosphorus is known to affect the productivity of aquatic systems; thus, the connection between changes in P loading (Fig. 1) and pro- ductivity is not surprising. However, a major advance in aquatic ecology was the quantification of the re- lationship between P and production across system2’. This quantification showed the overall importance of P in controlling production in lakes. Further, the use of quantitative models has been important for making predictions about human inputs due to changing land use and management practice?.

Testable predictions seem to become more difficult when com- plex interactions mediate the response to input variables32. These predictions will be aided by a better understanding of the bio- geochemical and trophic inter-

I Anoxic waters/sediments

Fig. 2. Indirect effects on the phosphorus cycle in aquatic systems by linked element cycles. The supply of P to sur- face waters is an important control of primary production in aquatic systems. Fe”compounds can reduce the supply of P to surface waters as these compounds trap P. Humans may be altering the P cycle indirectly through their influence on variables that control the effectiveness of the iron trap. Nitrate, sulfate and H’ can either alter the pool size of Fe” or modify the ability of Fe” to sorb P. The loading to aquatic systems of these three com- pounds has increased due to human activity in the water- sheds and airsheds. The + and - symbols indicate if the chemical compounds are thought to have a positive or negative influence on the cited processes.

actions in ecosystems, which in turn will allow us to tailor our predictions to classes of systems that might be most sensitive to different indirect perturbations. A simple example of this would be to select a subset of acid sensitive (low carbonate) lakes when looking for lake response to acid precipitatior?‘. A seemingly more complicated example is the

TREE vol. 8, no. 2, February 1993

Low r J buffering Geology

\ Mid/high buffering

Fig. 3. An example of classification of lakes to predict multiple stresses. Lakes were classified by geo- logic setting (buffering) and productivity to predict changes in P cycling due to the simultaneous inputs of H’ and sulfate. The buffering capacity of the system (related to geology of area) is an extremely important determinant of H+ sensitivity. The productivity of the system is a malor determinant of sensitivity to sul- fate as the amount of sulfate reduction to sulfide (Fig. 2) is dependent on the productivity of the system as well as the sulfate content74. The productivity of the system also affects the sensitivity to H+ as many of the processes that generate alkalinity within a lake are productivity dependent35. Finally, the sensi- tivity to sulfate inputs is likely to depend on the buffering capacity of the lake as iron sulfides are less likely to form (Fig. 2) under the low pH conditions that may be present in systems with low buffering capacity36. Owing to these interactions, low buffering systems with low productivity are likely to be H+- but not sulfur sensitive. Systems with high productivity and high buffering capacity are likely to be sul- phur but not H’-sensitive. The + and - symbols associated with the arrows indicate whether the ions are thought to have a positive or negative effect.

classification of lakes by their sensi- tivity to changes in the population of predatory fish (trophic cascade, Box 1). To date several classification schemes have been suggested32, but none has been widely accepted.

Redicting tCle impact of multiple disturbances

The prediction of impact of perturbation is even more difficult when a system receives multiple disturbances simultaneously. Human activity is, however, associated with multiple stresses. A predictive understanding of multiple stresses is necessary. An example of mul- tiple stresses is the simultaneous input of H+ and sulfate into aquatic systems. This simultaneous stress clearly happens as both inputs orig- inate largely from the burning of coal. The prediction of combined effects is seemingly difficult as H’ and sulfate have opposite effects on P cycling in aquatic systems’8r2’,22 (Figs 2 and 3). How will systems behave when both inputs occur? One possibility is that when H’ and sulfate enter a lake simultaneously the opposing effects of each ion will cause a net zero response22. 1 believe the general answer to how these effects combine is that the

54

response is mediated by several key characteristics of the system and, therefore, will vary among broad ‘classes’ of systems (Fig. 3).

Conclusions Human impacts are often in-

direct, being mediated through linked element cycles or trophic interactions. These indirect effects are in turn moderated by physical and geological characteristics of the system. A further complication is the fact that human disturbance is usually associated with multiple simultaneous impacts. Many of these complicated interactions are just now being elucidated for natu- ral systems. Intensive studies of biogeochemical and population interactions are necessary to more fully understand these complicated interactions. If we are to predict the consequence of human perturbation in more than a few well studied systems, however, we must try harder to use the knowledge from these intensive studies to make simplified models that are appli- cable to broad classes of systems.

Acknowledgements I thank 1.1. Cole, G.E. Likens and M.L. Pace

for their advice and assistance. This work

was supported in part by the A.W. Mellon Foundation and the National Science Foundation. This paper is a contribution to the program of the Institute of Ecosystems Studies.

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