STUDYING STRESS: THE IMPORTANCE OF ORGANISM-LEVEL RESPONSES

May 1999 431STRESS IN ECOLOGICAL SYSTEMS

431

Ecological Applications, 9(2), 1999, pp. 431–440q 1999 by the Ecological Society of America

STUDYING STRESS: THE IMPORTANCE OFORGANISM-LEVEL RESPONSES

LORRAINE MALTBY1

Department of Animal and Plant Sciences, The University of Sheffield, Sheffield, S10 2TN UK

Abstract. The importance of studying the effects of stress on individual organisms isaddressed by considering the use of individual-level information to: (1) elucidate the mech-anistic bases of interpopulation variation; (2) predict population-level effects; and (3) mon-itor stress in natural communities. Examples discussed include interpopulation variation inthe sensitivity of freshwater shrimps to zinc stress; the use of individual-based models topredict the effects of copper stress on earthworm populations; the use of single-species insitu assays to monitor pollution. It is contended that knowledge of organism-level responsesis essential for understanding how stressors cause adverse biological effects and the strat-egies adopted by organisms to tolerate stress. It is also contended that the effects of stressorson populations can be predicted from a knowledge of the effects of stressors on individualenergy budgets. Organism-level responses can be used to monitor stress in natural envi-ronments. In situ assays, based on the physiological energetics of ecologically relevantspecies, can provide sensitive and general stress indicators that are correlated with com-munity-level responses.

Key words: bioassays; energy budgets; Gammarus feeding rate; interpopulation variation, mech-anistic basis; physiological models, individual-based; population-level effects; stress monitoring usingbioassays; stress responses, individual and community level; zinc tolerance.

INTRODUCTION

Ecotoxicologists study the fate and effects of toxi-cants in ecosystems with the aim of understanding howtoxicants affect the structure and functioning of pop-ulations, communities, and ecosystems. Although theultimate level of concern may be populations, com-munities, or ecosystems, chemicals affect individualorganisms, and the consequences of stress may be man-ifested at all levels of biological organization. Thereis not a ‘‘right’’ level at which to study stress. Rather,different levels of organization provide informationthat, in combination, give insight into the effects ofstress, their mechanistic bases, and their ecological andevolutionary consequences. Studies of populations andcommunities can provide a description of the effectsof stress but do not, in themselves, provide informationon how effects are caused. Conversely, studies at themolecular and cellular level can provide detailed in-formation on how chemicals interact with target sitesbut provide little or no information on the conse-quences of these effects for higher levels of organi-zation. What is required is an integrated approach inwhich an understanding of the mechanistic bases of

Manuscript received 20 October 1997; accepted 7 July1998. For reprints of this Invited Feature, see footnote 1,page 429.

1 E-mail: [email protected]

stress responses in individuals is used to predict orinterpret their ecological consequences.

Here I will consider how understanding the physi-ological responses of individuals to stress can (1) pro-vide insight into the development of stress tolerance,and (2) be used to predict population-level effects.Knowledge of the mechanistic bases of stress toleranceis essential in order to understand why species differin their susceptibility to stressors and how populationscan persist in contaminated environments. Moreover,it is necessary to translate effects on individuals toeffects on populations in order to evaluate the ecolog-ical consequences of stress. Physiologically based in-dividual models combined with structured populationmodels provide a means of predicting population-leveleffects from information on individual organisms. Fi-nally, the application of individual-level physiologicalresponses to the study of stress effects on natural eco-systems will be discussed.

MECHANISTIC BASES FOR STRESS TOLERANCE

Numerous studies have demonstrated interpopula-tion variation in stress response; organisms from con-taminated populations are less sensitive than those fromuncontaminated populations (e.g., Klerks and Weis1987, Posthuma and Van Straalen 1993). Individualsfrom contaminated sites may be an order of magnitudeless sensitive than individuals of the same species fromuncontaminated sites (Table 1). In order to elucidate

432 INVITED FEATURE Ecological ApplicationsVol. 9, No. 2

TABLE 1. Sensitivity of invertebrates to chemical stress.

Species Stressor ResponseRelative

sensitivity Reference

PolychaetaNephthys hombergiNereis diversicolor

coppercopper

survivalsurvival

4.32.8

Bryan 1976Bryan 1976

OligochaetaLimnodrilus hoffmeisteri cadmium survival 4 Klerks and Levinton 1989

CrustaceaAsellus aquaticus lead

zincsurvivalsurvival

2.71.5

Fraser et al. 1978Naylor et al. 1990

Asellus meridianus lead

copper

survivalgrowthsurvivalgrowth

47.32.13.6

Brown 1976Brown 1976Brown 1976Brown 1976

Corophium volutatorGammarus pulex

copperzinccadmium

survivalsurvivalsurvival

1.657

Bryan 1976Roberts 1996Roberts 1996

Hyalella azteca low pH survival 2.8–12 France and Stokes1987

Porcellio scaber cadmium growth 6 Donker and Bogert1991

MolluscaOnychiurus armatus zinc/cadmium growth

reproduction22

Tranvik et al. 1993Tranvik et al. 1993

Orchesella cinctaScrobicularia plana

cadmiumcopper

growthsurvival

1.32

Posthuma 1990Bryan 1976

InsectaLucilia cuprina diflubenzuron survival 10 Kotze et al. 1997

Notes: For each species, the average response of individuals from a contaminated site is presented relative to the averageresponse of individuals from a reference site. Values of relative sensitivity .1 indicate that animals from the contaminatedpopulation are less sensitive to the stressor than those from the reference population.

the mechanisms resulting in differences in sensitivity,it is necessary to focus on the stress responses of in-dividuals. On exposure, chemicals enter individualsand may accumulate in body tissues. Interactions be-tween accumulated chemical and target moleculescause damage that may eventually result in disease,malfunction, and death. An accumulated chemical mayinduce defense mechanisms that reduce its effect ontarget tissues, and damage may be repaired, thus re-ducing the severity of the response.

There are several mechanisms by which the suscepti-bility of individuals to toxicants may be reduced. Theeffects of toxicant exposure may be reduced by decreaseduptake or increased depuration, enhanced detoxificationand sequestration processes, reduced target-site suscep-tibility, or improved repair processes (Brattsten et al.1986, Posthuma and Van Straalen 1993, McKenzie andBatterham 1994). For example, metal-tolerant poly-chaetes, Nereis diversicolor, were less permeable andhence absorbed less water-borne metal than nontolerantpolychaetes (Bryan and Hummerstone 1973). Increasedexcretion resulted in increased metal tolerance in themidge Chironomus riparius and the springtail Orchesellacincta (Van Straalen et al. 1987). Increased detoxificationhas been reported for the metal-tolerant oligochaete Lim-

nodrilus hoffmeisteri (Klerks and Levington 1989) andthe polychaetes Nereis diversicolor and Nephthys hom-bergi (Klerks and Weis 1987). Enhanced monooxygenaseactivity was correlated with diflubenzuron tolerance inthe Australian sheep blowfly Lucilia cuprina (Kotze etal. 1997) whereas organophosphate resistance in mos-quitoes (Culex pipiens) was due to elevated esterase Bproduction caused by gene amplification (Mouches et al.1986, Raymond et al. 1989). Increased storage capacityoccurs in metal-tolerant midges Chironomus riparius(Postma et al. 1996), and an example of reduced target-site sensitivity is provided by the leafhopper Nephotettixcincticeps. Strains of this species have acetylcholinester-ase that is insensitive to certain carbamate and organo-phosphate insecticides (Brattsten et al. 1986). Resistanceto cyclodien insecticides (e.g., dieldrin) provides anotherexample of target site modification. A point mutation,resulting in the substitution of serine for analine in aDrosophila GABA receptor, reduces binding by cyclodieninsecticides and confers resistance (ffrench-Constant etal. 1993).

Studies of metal tolerance have focused on uptake,excretion, detoxification, and sequestration. Populationvariation in response to metals may, however, be dueto other factors. Metal-tolerant Gammarus pulex, for


example, did not accumulate less metal than reference-site individuals and there was no evidence that indi-viduals from the two populations differed in metal de-toxification (i.e., metal-binding proteins) or sequestra-tion (i.e., storage in hepatopancreas) processes. Despitethe lack of interpopulation variation in metal accu-mulation and handling, individuals from the contami-nated site were five times less susceptible to zinc andseven times less susceptible to cadmium (Roberts1996). What is the mechanistic basis for this markeddifference in metal sensitivity? Gills are a major siteof aqueous metal uptake in crustaceans and are sen-sitive to metal-induced damage (e.g., Bubel, 1976,Couch 1977, Lawson et al. 1995) which may result inrespiratory and osmoregulatory impairment (Spicerand Weber 1991, 1992, Nonnotte et al. 1993). Inter-population differences in the effect of metal exposureon osmoregulation and susceptibility to respiratory im-pairment were observed with Gammarus pulex. Os-moregulation of individuals from the metal-contami-nated population was less sensitive to metal exposurethan that of individuals from the reference population(Spicer et al. 1998). Moreover, individuals from themetal-contaminated population were less sensitive tohypoxia and consequently less susceptible to metal-induced respiratory impairment (Roberts 1996). As os-moregulation was relatively insensitive to zinc expo-sure, it was concluded that the observed interpopula-tion difference in metal sensitivity was linked to dif-ferences in respiratory physiology. It appears thatreduced susceptibility to metal stress was a result ofphysiological adaptation rather than variation in ac-cumulation or defense mechanisms.

If studies of stress focused only on populations andcommunities, we would observe that organisms couldpersist in contaminated habitats but would not knowwhy. Only by studying the effects of stress on indi-viduals can we understand how organisms cope withstress and elucidate the mechanistic bases of stress tol-erance.

PREDICTING POPULATION-LEVEL EFFECTS

One of the major challenges facing ecotoxicologistsis evaluating the ecological relevance of individual-level effects. Does it matter that the behavior, physi-ology, or survival of individuals is impaired? What arethe consequences to the population if 50% of individ-uals are killed by a toxicant? These are difficult ques-tions to address without a detailed understanding of:(1) the individual and population-level consequencesof suborganism effects; (2) the processes regulatingpopulation size, and (3) the minimum viable populationsize. The first point can be addressed by understandingthe mechanisms by which changes in the performanceof individuals may impact the population to which theybelong.

As stated above, individuals may respond to a stres-

sor by inducing defense and repair processes. The pro-duction of detoxification enzymes (e.g., mixed-func-tion oxidases), metal-binding proteins (e.g., metallo-thionein), heat-shock proteins, and increased proteinturnover all require energy and may consequently resultin increased maintenance costs. Given that resourcesavailable to an individual are finite, increasing main-tenance costs will mean that fewer resources are avail-able for growth and reproduction. An example of thepossible trade-off between defense and reproduction isprovided by the work of Krebs and Loeschcke (1994).Heat-shock genes in Drosophila melanogaster were ac-tivated by exposing female flies to elevated tempera-tures for a short period of time. Flies pre-exposed toelevated temperatures were better able to survive sub-sequent thermal stress but were less fecund than controlflies. Krebs and Loeschcke (1994) argued that the re-duction in fecundity resulted from energy allocation toreproduction being reduced due to increased energyrequirements for heat-shock protein production. Morerecently, a negative association between heat-shockprotein concentration and survival in the absence ofstress has been reported (Krebs and Feder 1997), sug-gesting the presence of a trade-off between thermo-tolerance and performance.

The amount of energy available for growth and re-production can be estimated by measuring an organ-ism’s energy budget and calculating its ‘‘scope forgrowth.’’ Scope for growth (SfG) is defined as the dif-ference between energy absorbed from food and thatlost via excretion and metabolism (Warren and Davis1967), and is determined by measuring energy ingest-ed, egested, respired, and excreted. A positive SfG in-dicates that energy is available for production, while anegative SfG indicates that reserves must be used tomaintain the individual. Short-term measures of SfGcorrelate well with long-term measures of growth andreproduction (e.g., Bayne et al. 1985, Maltby and Nay-lor 1990, Maltby 1994). I have already postulated thatexposure to chemical stressors may increase energyexpenditure due to the costs of defense and repair pro-cesses. However, it is clear from the results of a numberof studies on a variety of species and stressors, thatexposure to toxicants generally results in a decrease infeeding and hence in energy acquisition (Table 2).Stress-induced reductions in energy acquisition oftenoccur at exposures lower than those known to affectmaintenance costs, measured as changes in oxygen up-take (e.g., Donkin and Widdows 1986, Maltby 1992).

Information on how toxicants affect energy budgetscan be incorporated into physiologically based modelsto predict the effects of toxicants on the growth, sur-vival, and reproduction of individuals (Calow and Sibly1990). Most work in this area has focused on the fresh-water cladoceran Daphnia; one of the earliest studieswas performed by Paloheimo et al. (1982). Data fromfeeding and assimilation experiments were combined


TABLE 2. Studies of stress-induced feeding inhibition in invertebrates.

Species Stressor Reference

CrustaceaCallinectes sapidus cadmium Guerin and Stickle 1995

Daphnia magna cadmiumvanadiumsodium bromide3,4-dichloroaniline

Allen et al. 1995Allen et al. 1995Allen et al. 1995Allen et al. 1995

Gammarus pulex low pHammonia3,4-dichloroanilinepentachlorophenolLindanezinccoppercadmium

Hargeby and Petersen 1988Maltby 1994Maltby et al. 1990aMaltby 1994Warwick (1997)Maltby et al. 1990aTattersfield 1993Stuhlbacher and Maltby 1992

Orchestia gammarellus copperzinc

Weeks 1993Weeks 1993

Oniscus asellus cobalt Drobne and Hopkin 1994

Porcellio scaber cadmiumcobaltzinc

Donker and Bogert 1991Drobne and Hopkin 1994Donker et al. 1996

MolluscaMytilus edulis Carbaryl

2,4-dichlorophenolDichlorvosEndrinFlucythrinateLindanepentachlorophenolPermethrin2,4,5-trichlorophenolTBT, DBT

Donkin et al. 1996, 1997Donkin et al. 1996, 1997Donkin et al. 1996, 1997Donkin et al. 1996, 1997Donkin et al. 1996, 1997Donkin et al. 1996, 1997Donkin et al. 1996, 1997Donkin et al. 1996, 1997Donkin et al. 1996, 1997Widdows and Page 1993

Helix aspersa AminocarbAzinphosmethylCarbarylFenitrothionmethyl parathionParaquatTrichlorfon

Schuytema et al. 1994Schuytema et al. 1994Schuytema et al. 1994Schuytema et al. 1994Schuytema et al. 1994Schuytema et al. 1994Schuytema et al. 1994

Brotia hainanensis copperlow pHcadmium

Lai and Lam 1994Lai and Lam 1994Lam 1996

Thais lapillus oil Stickle 1985

with literature values on respiration rate to predictgrowth and reproduction. Predicted and observedgrowth curves were very similar (Fig. 1a), but the mod-el underestimated reproduction by as much as 43%(Fig. 1b). One possible reason for this discrepancy be-tween predicted and actual reproduction is that energybudget measurements, including SfG, provide an in-stantaneous measure of the energy available for pro-duction and do not take into account production fromenergy reserves. Subsequent energy budget modelshave incorporated a storage component, although theprecise allocation rules vary between models. In somemodels, assimilate enters the blood/reserve compart-ment from which it is allocated to reproduction orgrowth plus maintenance (Kooijman 1986). In others,assimilate is allocated immediately to reproduction or

to growth (including storage) and maintenance (Gurneyet al. 1990, McCauley et al. 1990).

By combining physiologically based individual mod-els with demographic population models it is possible,in principle, to use information based on studies ofindividuals to predict population-level responses. Twotypes of models have been used to predict populationdynamics from individual responses: those that use av-erage values and treat all individuals as if they werethe same, and those that predict population dynamicsby specifying the variation in responses of individualmembers of that population (Metz et al. 1988, Caswell1989, DeAngelis and Gross 1992). Whereas the formeris less data intensive and computationally less de-manding, the latter is more realistic as it incorporatesintrapopulation variation. The data requirements of in-


FIG. 1. Predicted (open symbols) and observed (solidsymbols) (a) growth and (b) reproduction of Daphnia pulexthat were fed Chlamydomonas reinhardtii (104 cells/mL) andmaintained at 188C. Predictions are based on an energy budgetmodel (Paloheimo et al. 1982).

dividual-based models may be reduced by groupingindividuals into distinct age or size classes or devel-opmental stages (e.g., embryo, juvenile, adult), and de-veloping an age-, size-, or stage-structured populationmodel. Kooijman and Metz (1984), for example, usedan age-structured model to predict the dynamics ofchemically stressed Daphnia populations from infor-mation on the effect of toxicants on individual energybudgets. This energy budget approach has since beendeveloped and applied to a range of organisms (Kooij-man 1993).

Klok and de Roos (1996) used an individual-basedmodel (Kooijman and Metz 1984) to predict the effectof copper-induced changes in feeding and metabolismon the growth and reproduction of the worm Lumbricusrubellus. The model assumes that a fixed proportion ofenergy is spent on maintenance and growth, the re-mainder being allocated to reproduction. It also as-sumes that energy requirements for maintenance havepriority over those for growth. Klok and de Roos (1996)investigated two different effects of toxic stress on en-ergy budgets: (1) a decrease in energy assimilated, and(2) an increase in maintenance costs. The model pre-

dicted that, over the concentration range tested (13–362 mg Cu/kg), reproduction would be reduced by atoxicant-induced reduction in assimilation, but not bya toxicant-induced increase in maintenance costs. Bothscenarios resulted in a decrease in individual growth.Predictions of individual performance were translatedinto population-level consequences using a size-struc-tured matrix model. With both scenarios, populationsexposed to the highest test concentration (i.e., 362 mg/kg) would become extinct, even though there was noeffect of increased maintenance costs on reproduction.The reason for the population decline was the severereduction in individual growth resulting from eitherreduced energy assimilation or increased maintenancecosts. The impairment of growth was such that animalswould not attain reproductive size and therefore be in-capable of reproducing. Using this approach, Klok andde Roos (1996) predicted copper concentrations at acritical level (i.e., population growth rate equals zero)of between 200 and 300 mg Cu/kg, which correspondedwell to the observed steep decline in the size of fieldpopulations at copper concentrations exceeding 200 mgCu/kg (Ma 1988).

Individual-based models have been developed andtested using carefully controlled laboratory systems.However, the dynamics of natural populations are de-termined, not only by the properties of individuals mak-ing up that population, but also by a variety of bioticand abiotic factors, the interactions between which maybe altered by the presence of a chemical stressor. Forexample, chemical stressors are known to alter pred-ator–prey and competitive interactions (e.g., Clementset al. 1989, Ferrando et al. 1993), which may confoundor mask the direct effects of stressors on individuals.Another limitation of energy budget models is that,whereas they can provide a prediction of mortality dueto starvation, they do not, in themselves, predict mor-tality due to the direct effects of toxicants with specificmodes of action. These data have to be derived fromdose-response relationships determined experimental-ly. Furthermore, assessing the importance of toxicant-induced mortality to overall population size and struc-ture is far from straightforward. If a population is understrong density-dependent control, the loss of 50% ofjuveniles due to the presence of a toxicant may havelittle effect on overall population dynamics, but wouldaffect the genetic structure of the population. Thesecriticisms are not, however, meant to constitute an ar-gument for abandoning an individual-based approach.Rather, they are an appreciation that, whereas infor-mation about the effects of toxicants on individualsmay provide insight into the potential effect of stressorson populations and the communities to which they be-long, the actual effects observed will be a function ofa number of interacting factors.


TABLE 3. Use of scope for growth or feeding inhibition to monitor pollution.

Species Pollution Response† Reference

MolluscaAnadara granosa industrial SfG Din and Ahamdad 1995Mytilus edulis oil

hydrocarbons, TBTand organochlorines

SfGSfG

Widdows et al. 1995Widdows et al. 1995

industrial SfG Roddie et al. 1996Mytilus

galloprovincialishydrocarbons, PCBs,

pesticidesSfG Widdows et al. 1996

CrustaceaGammarus pulex Malathion

Carbofuranorganic enrichmentmetalshydrocarbonslow pH, aluminumzinc

feedingfeedingfeedingfeedingfeedingfeedingfeeding

Crane et al. 1995Matthiessen et al. 1995Veerasingham and Crane 1992Crane and Maltby 1991Forrow 1995McCahon et al. 1989Roddie et al. 1992

† SfG 5 scope for growth (see Predicting population-level effects).

APPLICATION: MONITORING STRESS

USING INDIVIDUALS

Measures of community structure have traditionallybeen used to monitor the effects of stress on aquaticecosystems (Rosenberg and Resh 1993). Diversity in-dices (Washington 1984) provide good measures of thecommunity structure but take no account of speciesidentity. Biotic indices incorporate taxa-specific infor-mation by weighting taxa according to their sensitivityto pollution (Metcalfe 1989). Diversity indices detectchanges in the number of species present and theirrelative abundance, and biotic indices detect changesin the number of species or families present in a com-munity. Multivariate ordination and classification tech-niques can be used to compare communities, and togroup together sites with similar communities (e.g.,Gauch, 1982, Clarke and Warwick 1994, Manly 1994).

Many factors, both natural and anthropogenic, influ-ence the structure of communities. If differences incommunity structure are detected, it is not possible toestablish to what extent the changes observed arecaused by the stressor being investigated. Nor is it pos-sible to establish how the stressor results in changesin community structure. Species may be reduced inabundance either because of the direct effects of thestressor or because of effects on competitors, predators,or prey (DeAngelis 1996). These questions can onlybe addressed by knowing how stressors affect individ-ual organisms and understanding the consequences ofthese effects for the populations and communities towhich the individuals belong.

Community-level stress measures, such as biotic anddiversity indices or multivariate analyses, provide use-ful and potentially sensitive descriptions of communitystructure, but they are insensitive to sublethal levels ofstress (Gray et al. 1990, Dawson-Shepherd et al. 1992).In contrast, individual-level measures of stress can beused to detect sublethal effects, and have the added

advantage of short response times and the potential toprovide information on causal agents. Single-speciesin situ assays provide useful tools for studying stressin terrestrial, marine, and freshwater ecosystems. Inaquatic systems, caged fish have been used to detectthe presence of oestrogenic chemicals (Purdom et al.1994), mussels have been used to detect pollution fromdump sites, oil terminals, and industrial discharges(Roddie et al. 1996) and crustaceans have been usedto study the impact of acid pulses (McCahon et al.1989), pesticides (Matthiessen et al. 1995), and met-alliferous discharges (Crane and Maltby 1991).

Single-species in situ assays use resident or trans-planted organisms, and measurement endpoints arebased on lethal or sublethal responses (Crane et al.1996). Whole-organism endpoints such as death, mor-phological deformities, and growth, provide generalmeasures of stress and can be used to detect changesin environmental quality. Molecular-level responsesmay be more toxicant specific and have the potentialto be used as diagnostic tools. For example, inhibitionof acetylcholinesterase has been used as an indicatorof organophosphate and carbamate pesticide exposure,and inhibition of aminolevulinic acid dehydratase hasbeen used as a indicator of lead exposure (Peakall1992).

A key question to be addressed with any single-spe-cies approach is: Which species should be used? Stan-dard laboratory species include freshwater cladocerans(e.g., Daphnia magna), marine copepods (e.g., Acartiatonsa), marine bivalves (e.g., Crassostrea gigas, My-tilus edulis), freshwater and marine fish (e.g., Onco-rhynchus mykiss, Pimephales promelas, Cyprinodonvariegatus), earthworms (e.g., Eisenia foetida), andbirds (e.g., Anas platyrhynchos, Colinus virginianus)(Calow 1998). These standard species may not, how-ever, be the best species to use in in situ studies. Thechoice of test species should be driven by the precise


question being addressed, but, in general, test speciesshould be sensitive and ecologically relevant. The se-lection of test species therefore requires some under-standing of the system to be monitored and the relativesensitivity of the species present. Community-basedstudies could be used to inform this selection processby identifying sensitive species in relevant habitats andassemblages.

By selecting ecologically relevant species and end-points it is possible to use the results of single-speciesin situ assays as short-term predictors of more long-term effects on populations and communities. The am-phipod Gammarus pulex is an important detritivore inmany freshwater ecosystems and is used in a 6-d insitu feeding rate assay (Maltby et al. 1990b, Crane andMaltby 1991). The rate of detritus processing instreams is dependent upon the feeding rate of detriti-vores such as Gammarus pulex (Maltby 1996). Severallaboratory and field studies have demonstrated thatGammarus feeding rate is depressed by exposure to avariety of chemical stressors (Maltby 1994). Studies inartificial streams have demonstrated a positive corre-lation between Gammarus feeding rate, measured in a6-d assay, and the processing of leaf material by benthicmacroinvertebrates (Tattersfield 1993). More recentstudies have demonstrated a similar relationship be-tween Gammarus feeding rate and leaf processing innatural streams (S. Clayton, unpublished data). More-over, as discussed earlier, toxicant-induced changes infeeding rate, and consequently SfG, are indicative oflonger term changes in growth and reproduction andcan be used to predict population-level effects. Thefeeding rate and SfG of molluscs and crustaceans havebeen used to monitor different types of aquatic pol-lution (Table 3).

CONCLUSIONS

1) Organism-level responses are essential for elu-cidating the mechanistic bases of interpopulation dif-ference in stress tolerance. Commonly reported mech-anisms of reduced susceptibility are: decreased uptake;increased excretion, detoxification, sequestration, re-pair; decreased target-site sensitivity. Susceptibilitymay also be reduced by adaptation, genetic or phe-notypic, of physiological processes. Interpopulationdifferences in the susceptibility of G. pulex to zinc wereassociated with differences in respiratory physiology.

2) Organism-level responses can be used to predictthe effect of stress on populations. Individuals exposedto stress have reduced energy intake and possibly en-hance energy expenditure on defense and repair pro-cesses. Individual-based energy budget models andstructured population models have been used to trans-late stress-induced changes in energy acquisition andallocation to population-level effects.

3) Organism-level responses provide a sensitive andgeneral stress indicator that is correlated with com-

munity-level effects. The feeding rate of G. pulex isinhibited by a variety of stressors and is positivelycorrelated with leaf processing by benthic communi-ties. The in situ Gammarus feeding rate assay can beused to indicate population- and community-level ef-fects. The use of in situ assays that combine generalstress responses and toxicant-specific molecular re-sponses can provide ecologically relevant, sensitive,and diagnostic monitoring tools.

LITERATURE CITED

Allen, Y., P. Calow, and D. Baird. 1995. A mechanistic modelof contaminant-induced feeding inhibition in Daphnia mag-na. Environmental Toxicology and Chemistry 14:1625–1630.

Bayne, B. L., D. A. Brown, K. Burns, D. R. Dixon, A. Iva-novici, D. R. Livingstone, D. M. Lowe, M. N. Moore, A.R. D. Stebbing, and J. Widdows. 1985. The effects of stressand pollution on marine animals. Praeger, New York, NewYork, USA.

Brattsten, L. B., C. W. Holyoke, J. R. Leeper, and K. F. Raffa.1986. Insecticide resistance: challenge to pest managementand basic research. Science 231:1255–1260.

Brown, B. E. 1976. Observations on the tolerance of theisopod Asellus meridianus Rac. to copper and lead. WaterResearch 10:555–559.

Bryan, G. W. 1976. Some aspects of heavy metal tolerancein aquatic organisms. Pages 7–34 in A. P. M. Lockwood,editor. Effects of pollutants on aquatic organisms. Cam-bridge University Press, Cambridge, UK.

Bryan, G. W., and L. G. Hummerstone. 1973. Adaptation ofthe polychaete Nereis diversicolor to estuarine sedimentscontaining high concentrations of zinc and cadmium. Jour-nal of the Marine Biological Association of the UK 53:839–857.

Bubel, A. 1976. Histological and electron microscopical ob-servations on the effects of different salinities and heavymetal ions, on the gills of Jaera normanni (Rathke) (Crus-tacea: Isopoda). Cell and Tissue Research 167:65–95.

Calow, P. 1998. Handbook of ecotoxicology. Blackwell Sci-ence, Oxford, UK.

Calow, P., and R. M. Sibly. 1990. A physiological basis ofpopulation processes: ecotoxicological implications. Func-tional Ecology 4:283–288.

Caswell, H. 1989. Matrix population models: construction,analysis and interpretation. Sinauer Associates, Sunder-land, Massachusetts, USA.

Clarke, K. R., and R. M. Warwick. 1994. Change in marinecommunities: an approach to statistical analysis and inter-pretation. Natural Environment Research Council, London,UK.

Clements, W. H., D. S. Cherry, and J. Cairns, Jr. 1989. Theinfluence of copper exposure on predator–prey interactionsin aquatic insect communities. Freshwater Biology 21:483–488.

Couch, J. A. 1977. Ultrastructure study of lesions in gills ofa marine shrimp exposed to cadmium. Journal of Inverte-brate Pathology 29:267–288.

Crane, M., P. Delaney, P. Parker, C. Walker, and S. Watson.1995. The effect of Malathion 60 on Gammarus pulex (L.)below watercress beds. Environmental Toxicology andChemistry 14:1181–1188.

Crane, M., I. Johnson, and L. Maltby. 1996. In situ assaysfor monitoring the toxic impacts of waste in rivers. Pages116–124 in J. F. Tapp, S. M. Hunt, and J. R. Wharfe, editors.Toxic impacts of wastes on the aquatic environment. RoyalSociety of Chemistry, Cambridge, UK.

Crane, M., and L. Maltby. 1991. Lethal and sublethal re-


sponses of Gammarus pulex to stress: sensitivity and vari-ation in an in situ assay. Environmental Toxicology andChemistry 10:1331–1339.

Dawson-Shepherd, A., R. M. Warwick, K. R. Clarke, and B.E. Brown. 1992. An analysis of fish community responsesto coral mining in the Maldives. Environmental Biology ofFishes 33:367–380

DeAngelis, D. L. 1996. Indirect effects: concepts and ap-proaches from ecological theory. Pages 25–41 in D. J.Baird, L. Maltby, P. W. Greig-Smith, and P. E. T. Douben,editors. Ecotoxicology: ecological dimensions. Chapmanand Hall, London, UK.

DeAngelis, D. L., and J. G. Gross. 1992. Individual-basedmodels and approaches in ecology. Populations, commu-nities and ecosystems. Chapman and Hall, London, UK.

Din, Z. B., and A. Ahamdad. 1995. Changes in the scopefor growth of blood cockles (Anadara granosa) exposed toindustrial discharge. Marine Pollution Bulletin 31:406–410.

Donker, M. H., and C. G. Bogert. 1991. Adaptation to cad-mium in three populations of the isopod Porcellio scaber.Comparative Biochemistry and Physiology 100C:143–146.

Donker, M. H., M. H. Raedecker, and N. M. Van Straalen.1996. The role of zinc regulation in the zinc tolerancemechanism of the terrestrial isopod Porcellio scaber. Jour-nal of Applied Ecology 33:955–964.

Donkin, P., and J. Widdows. 1986. Scope for growth as ameasurement of environmental stress and pollution. Studieswith bivalve molluscs. Philosophical Transactions of theRoyal Society of London, B. Biological Science 286:563–581.

Donkin, P., J. Widdows, S. V. Evans, F. J. Staff, and T. Yan.1997. Effect of neurotoxic pesticides on the feeding rateof marine mussels (Mytilus edulis). Pesticide Science 49:196–209.

Donkin, P., J. Widdows, D. M. Lowe, M. E. Donkin, and D.N. Price. 1996. Sub-lethal effects of waste on marine or-ganisms: responses measured at the whole-organism level.Pages 138–145 in J. F. Tapp, S. M. Hunt, and J. R. Wharfe,editors. Toxic impacts of wastes on the aquatic environ-ment. Royal Society of Chemistry, Cambridge, UK.

Drobne, D., and S. P. Hopkin. 1994. Ecotoxicological lab-oratory test for assessing the effects of chemicals on ter-restrial isopods. Bulletin of Environmental Contaminationand Toxicology 53:390–397.

Ferrando, M. D., C. Janssen, E. Andreu, and G. Persoone.1993. Ecotoxicological studies with the freshwater rotiferBranchionus calyciflorus. Resource competition betweenrotifers and daphnids under toxic stress. Science of theTotal Environment Supplement 1993:1059–1069.

ffrench-Constant, R. H., T. A. Rocheleau, J. C. Steichen, andA. E. Chalmers. 1993. A point mutation in a DrosophilaGABA receptor confers insecticide resistance. Nature 363:449–451.

Forrow, D. M. 1995. Toxicological assessment of the effectof motorway runoff on stream macroinvertebrate commu-nity structure and function. Dissertation. University ofSheffield, Sheffield, UK.

France, R. L., and P. M. Stokes. 1987. Life stage and pop-ulation variation in resistance and tolerance of Hyalellaazteca (Amphipoda) to low pH. Canadian Journal of Fish-eries and Aquatic Sciences 44:1102–1111.

Fraser, J., D. T. Parkin, and E. Verspoor. 1978. Tolerance tolead in the freshwater isopod Aseltus aquaticus. Water Re-search 12:637–641.

Gauch, H. G. 1982. Multivariate analysis in community ecol-ogy. Cambridge University Press, Cambridge, UK.

Gray, J. S., K. R. Clarke, R. M. Warwick, and G. Hobbs.1990. Detection of initial effects of pollution on marine

benthos: an example from the Ekofisk and Eldfisk oilfields,North Sea. Marine Ecology Progress Series 66:285–299.

Guerin, J. L., and W. B. Stickle. 1995. Effects of cadmiumon survival, osmoregulatory ability and bioenergetics ofthe juvenile blue crabs, Callinectes sapidus at different sa-linities. Marine Environmental Research 40:227–246.

Gurney, W. S. C., E. McCauley, R. M. Nisbet, and W. W.Murdoch. 1990. The physiological ecology of Daphnia: adynamic model of growth and reproduction. Ecology 71:716–732.

Hargeby, A., and R. C. Petersen. 1988. Effects of low pHand humus on the survivorship, growth and feeding ofGammarus pulex. Freshwater Biology 19:235–247.

Klerks, P. L., and J. S. Levinton. 1989. Rapid evolution ofmetal resistance in a benthic oligochaete inhabiting a metal-polluted site. Biological Bulletin 176:135–141.

Klerks, P. L., and J. S. Weis. 1987. Genetic adaptation toheavy metals in aquatic organisms: a review. Environmen-tal Pollution 45:173–205.

Klok, C., and A. M. de Roos. 1996. Population level con-sequences of toxicological influences on individual growthand reproduction in Lumbricus rubellus (Lumbricidae, Oli-gochaeta). Ecotoxicology and Environmental Safety 33:118–127.

Kooijman, S. A. L. M. 1986. Population dynamics on thebasis of budgets. Pages 266–297 in J. A. J. Metz and O.Diekman, editors. The dynamics of physiologically struc-tured populations. Springer-Verlag, Berlin, Germany.

. 1993. Dynamic energy budgets in biological sys-tems. Cambridge University Press, Cambridge, UK.

Kooijman, S. A. L. M., and J. A. J. Metz. 1984. On thedynamics of chemically stressed populations: the deductionof population consequences from effects on individuals.Ecotoxicology and Environmental Safety 8:254–274.

Kotze, A. C., N. Sales, and I. M. Barchia. 1997. Difluben-zuron tolerance associated with monooxygenase activity infield strain larvae of the Australian sheep blowfly (Diptera:Calliphoridae) Journal of Economic Entomology 90:15–20.

Krebs, R. A., and M. E. Feder. 1997. Natural variation inthe expression of the heat-shock protein HSP70 in a pop-ulation of Drosophila melanogaster and its correlation withtolerance of ecologically relevant thermal stress. Evolution51:173–179.

Krebs, R. A., and V. Loeschcke. 1994. Costs and benefits ofactivation of the heat-shock response in Drosophila mel-anogaster. Functional Ecology 8:730–737.

Lai, P. C. C., and P. K. S. Lam. 1994. Scope for growth ina tropical freshwater snail, Brotia hainanensis: implicationsfor monitoring sublethal toxic stressors. Internationale Ver-einisung fur theoretische und angewandte Limnologie, Mit-teilungen 24:315–320.

Lam, P. K. S. 1996. Sublethal effects of cadmium on theenergetics of a tropical freshwater snail, Brotia hainanensis(Brot, 1872). Environmental Toxicology and Water Quality11:345–349.

Lawson, S. L., M. B. Jones, and R. M. Moate. 1995. Effectof copper on the ultrastructure of the gill epithelium ofCarcinus maenas (Decapoda: Brachyura). Marine PollutionBulletin 31:187–196.

Ma, W. 1988. Toxicity of copper to lumbricid earthwormsin sandy agricultural soils amended with Cu-enriched or-ganic waste materials. Ecological Bulletin-NFR 39:53–56.

Maltby, L. 1992. The use of the physiological energetics ofGammarus pulex to assess toxicity: a study using artificialstreams. Environmental Toxicology and Chemistry 11:79–85.

. 1994. Stress, shredders and streams: using Gam-marus energetics to assess water quality. Pages 98–110 inD. W. Sutcliffe, editor. Water quality and stress indicatorsin marine and freshwater systems: linking levels of organ-


isation. Freshwater Biological Association, Ambleside,Cumbria, U.K.

. 1996. Detritus processing. Pages 145–167 in G. Pettsand P. Calow, editors. River biota. Diversity and dynamics.Blackwell Science, Oxford, UK.

Maltby, L., and C. Naylor. 1990. Preliminary observationson the ecological relevance of the Gammarus ‘scope forgrowth’ assay: effect of zinc on reproduction. FunctionalEcology 4:393–397.

Maltby, L., C. Naylor, and P. Calow. 1990a. Effect of stresson a freshwater benthic detritivore: scope for growth inGammarus pulex. Ecotoxicology and Environmental Safety19:285–291.

Maltby, L., C. Naylor, and P. Calow. 1990b. Field deploy-ment of a scope for growth assay involving Gammaruspulex, a freshwater benthic invertebrate. Ecotoxicology andEnvironmental Safety 19:292–300.

Manly, B. F. J. 1994. Multivariate statistical methods. Aprimer. Second edition. Chapman and Hall, London, UK.

Matthiessen, P., D. Sheahan, R. Harrison, M. Kirby, R. Ry-croft, A. Turnbull, C. Volkner, and R. Williams. 1995. Useof a Gammarus pulex bioassay to measure the effects oftransient carbofuran runoff from farmland. Ecotoxicologyand Environmental Safety 30:111–119.

McCahon, C. P., A. F. Brown, M. J. Poulton, and D. Pascoe.1989. Effects of acid, aluminium and lime additions onfish and invertebrates in a chronically acidic Welsh stream.Water, Air and Soil Pollution 45:345–359.

McCauley, E., W. W. Murdoch, R. M. Nisbet, and W. S. C.Gurney. 1990. The physiological ecology of Daphnia: de-velopment of a model of growth and reproduction. Ecology71:703–715.

McKenzie, J. A., and P. Batterham. 1994. The genetic, mo-lecular and phenotypic consequences of selection for in-secticide resistance. Trends in Ecology and Evolution 9:166–169.

Metcalfe, J. L. 1989. Biological water quality assessment ofrunning waters based on macroinvertebrate communities:history and present status in Europe. Environmental Pol-lution 60:101–139.

Metz, J. A. J., A. M. de Roos, and F. van den Bosch. 1988.Population models incorporating physiological structure: aquick survey of the basic concepts and an application tosize-structured population dynamics in waterfleas. Pages106–126 in B. Ebenman and L. Persson, editors. Size-struc-tured populations. Springer-Verlag, Berlin, Germany.

Mouches, C., N. Pasteur, J. B. Berge, O. Hyrien, M. Raymond,R. Rober de Saint Vincent, M. de Silvestri, and G. P. Geor-ghiou. 1986. Amplification of an esterase gene is respon-sible for insecticide resistance in a California Culex mos-quito. Science 233:778–780.

Naylor, C., L. Pindar, and P. Calow. 1990. Inter- and intra-specific variation in sensitivity to toxins; the effects ofacidity and zinc on the freshwater crustaceans Asellus aqua-ticus (L.) and Gammarus pulex (L.). Water Research 24:757–762.

Nonnotte, L., F. Boitel, and J. P. Truchot. 1993. Waterbornecopper causes gill damage and haemolymph hypoxia in theshore crab Carcinus maenas. Canadian Journal of Zoology71:1569–1576.

Paloheimo, J. E., S. J. Crabtree, and W. D. Taylor. 1982.Growth model of Daphnia. Canadian Journal of Fisheriesand Aquatic Sciences 39:598–606.

Peakall, D. 1992. Animal biomarkers as pollution indicators.Chapman and Hall, London, UK.

Posthuma, L. 1990. Genetic differentiation between popu-lations of Orchesella cincta (Collembola) from heavy metalcontaminated sites. Journal of Applied Ecology 27:609–622.

Posthuma, L., and N. M. Van Straalen. 1993. Heavy-metal

adaptation in terrestrial invertebrates: a review of occur-rence, genetics, physiology and ecological consequences.Comparative Biochemistry and Physiology 106C:11–38.

Postma, J. F., P. Van Nugteren, and M. B. Buckert De Jong.1996. Increased cadmium excretion in metal-adapted pop-ulations of the midge Chironomus riparius (Diptera). En-vironmental Toxicology and Chemistry 15:332–339.

Purdom, C. E., P. A. Hardiman, V. J. Bye, N. C. Eno, C. R.Tyler, and J. P. Sumpter. 1994. Estrogenic effects of ef-fluents from sewage treatment works. Chemistry and Ecol-ogy 8:275–285.

Raymond, M., V. Beyssat-Arnaouty, N. Sivasubramanian, C.Mouches, G. P. Georghiou, and N. Pasteur. 1989. Ampli-fication of various esterase B’s responsible for organo-phosphate resistance in Culex mosquitoes. Biochemical Ge-netics 27:417–423.

Roberts, D. A. 1996. Variation in sensitivity to metals intwo freshwater crustaceans, Asellus aquaticus and Gam-marus pulex: observations and mechanisms. Dissertation.University of Sheffield, Sheffield, UK.

Roddie, B., T. Kedwards, and M. Crane. 1992. Potentialimpact of watercress farm discharges on the freshwateramphipod, Gammarus pulex L. Bulletin of EnvironmentalContamination and Toxicology 48:63–69.

Roddie, B., C. J. Redshaw, and S. Nixon, S. 1996. Sub-lethalbiological effects monitoring using the common mussel(Mytilus edulis): comparison of laboratory and in situ ef-fects of an industrial effluent discharge. Pages 125–137 inJ. F. Tapp, S. M. Hunt, and J. R. Wharfe, editors. Toxicimpacts of wastes on the aquatic environment. Royal So-ciety of Chemistry, Cambridge, UK.

Rosenberg, D. M., and V. H. Resh. 1993. Freshwater bio-monitoring and benthic macroinvertebrates. Chapman andHall, New York, New York, USA.

Schuytema, G. S., A. V. Nebeker, and W. L. Griffis. 1994.Effects of dietary exposure to forest pesticides on the browngarden snail Helix aspersa. Archives of EnvironmentalContamination and Toxicology 26:23–28.

Spicer, J. I., D. Morritt, and L. Maltby. 1998. Effect of water-borne zinc on osmoregulation in the freshwater amphipodGammarus pulex (L.) from populations that differ in theirsensitivity to metal stress. Functional Ecology 12:242–247.

Spicer, J. I., and R. E. Weber. 1991. Respiratory impairmentin crustaceans and molluscs due to exposure to heavy met-als. Comparative Biochemistry and Physiology 100C:339–342.

Spicer, J. I., and R. E. Weber. 1992. Respiratory impairmentby water-borne copper and zinc in the edible crab Cancerpagurus (L.) (Crustacea: Decapoda) during hypoxic ex-posure. Marine Biology 112:429–435.

Stickle, W. B. 1985. Effects of environmental factor gradi-ents on scope for growth in several species of carnivorousmarine invertebrates. Pages 601–616 in J. S. Gray and M.E. Christiansen, editors. Marine biology of polar regionsand effects of stress on marine organisms. John Wiley andSons, New York, New York, USA.

Stuhlbacher, A., and L. Maltby. 1992. Cadmium resistancein Gammarus pulex (L.) Archives of Environmental Con-tamination and Toxicology 22:319–324.

Tattersfield, L. J. 1993. Effects of copper on the energy bud-get of a stream detritivore: validation and ecological rel-evance. Dissertation. University of Sheffield, Sheffield,UK.

Tranvik, L., G. Bengtsson, and S. Rundgren. 1993. Relativeabundance and resistance traits of two Collembola speciesunder metal stress. Journal of Applied Ecology 30:43–52.

Van Straalen, N. M., T. B. A. Burghouts, M. J. Doornhof, G.M. Groot, M. P. M. Janssen, E. N. G. Joosse, J. H. VanMeerendonk, J. P. J. J. Theeuwen, H. A. Verhoef, and H.R. Zoomer. 1987. Efficiency of lead and cadmium excre-


tion in populations of Orchesella cincta (Collembola) fromvarious contaminated forest soils. Journal of Applied Ecol-ogy 24:953–968.

Veerasingham, M., and M. Crane. 1992. Impact of farmwaste on freshwater invertebrate abundance and the feedingrate of Gammarus pulex L. Chemosphere 25:869–874.

Warren, C. E., and G. E. Davis. 1967. Laboratory studies onthe feeding, bioenergetics and growth of fish. Pages 175–214 in S. D. Gerking, editor. The biological basis of fresh-water fish production. Blackwell Scientific, Oxford, UK.

Warwick, O. W. T. 1997. The use of a biomarker to assessthe effect of xenobiotic exposure on the freshwater inver-tebrate Gammarus pulex. Dissertation. University of Shef-field, Sheffield, UK.

Washington, H. G. 1984. Diversity, biotic and similarity in-dices. A review with special relevance to aquatic ecosys-tems. Water Research 18:653–692.

Weeks, J. M. 1993. Effects of dietary copper and zinc con-centrations on feeding rates of two species of talitrid am-

phipods (Crustacea). Bulletin Environmental Contamina-tion and Toxicology 50:883–890.

Widdows, J., P. Donkin, M. D. Brinsley, S. V. Evans, P. N.Salkeld, A. Franklin, R. J. Law, and M. J. Waldock. 1995.Scope for growth and contaminant levels in North Sea mus-sels, Mytilus edulis. Marine Ecology-Progress Series 127:131–148.

Widdows, J., P. Donkin, S. V. Evans, D. S. Page, and P. N.Salkeld. 1994. Sublethal biological effects and chemicalcontaminants monitoring of Sullom Voe (Shetland) usingmussels (Mytilus edulis). Proceedings of the Royal Societyof Edinburgh Section B 103:99–112.

Widdows, J., C. Nasci, and V. U. Fossato. 1996. Effects ofpollution on the scope for growth of mussels (Mytilus gal-loprovincialis) from the Venice Lagoon, Italy. Marine En-vironmental Research 43:69–79.

Widdows, J., and D. S. Page. 1993. Effects of tributyltin anddibutyltin on the physiological energetics of the mussel,Mytilus edulis. Marine Environmental Research 35:233–249.

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