Document J3KBGOeDV9RG5Ngebv2Bb6v8Z

Environmental health Perspectives VoL 49. pp. 247-260. 1983 URL 03673 Developing Standards for Environmental Toxicants: The Need to Consider Abiotic Environmental Factors and MicrobeMediated Ecologic Processes by H. Babich* and G. Stotzky* This article raneelc and diKWtn two novel iipecta (or the formulation of standards for environ mental toxicants. First, uniform national atandards for each pollutant will be underprolective for some ecosystems and overprolective for others, inasmuch as the toxicity of a pollutant to the indige nous biota is dependent on the physicochemical properties of the recipient environment. As the number of ehemieals that need regulation Is immense and as microbes appear to respond similarly to pollutant-abiotic factor interactions as do plants and animals, it is suggested that microbial as says be used initially to identify those abiotic factors that most influence the toxicity of specific pol lutants. Thereafter, additional studies using plants and animals can focus on these pollut ant-abiotic factor interactions, and more meaningful standards can then be formulated more rapid ly and inexpensively. Second, it is suggested that the response to pollutants of microbe-mediated ecologic processes be used to quantitate the sensitivity of different ecosystems to various toxicants. Such a quantification, expressed in terms of an "ecological dose 50%'* (EcDk could be easily incor porated into the methodologies currently used to set water quality criteria and would also be appli cable to setting criteria for terrestrial ecosystems. Introduction Through a variety of Federal statutes, including the Clean Air Act (CAA) of 1970, the Clean Water Act (CWA) as amended in 1977, the Federal Insecti cide, Fungicide, and Rodenticide Act (FIFRA) of 1972, the Federal Water Pollution Control Act (FWPCA) of 1972, the Resource Conservation and Recovery Act (RCRA) of 1976, and the Toxic Sub stances Control Act (TSCA) of 1976, the United States Environmental Protection Agency (EPA) is charged with protecting the health and welfare of human beings and of the environment from harmful exposures to toxic agents Cl I. For example, as required by the CAA, EPA set national primary standards to protect human health against toxic levels of atmospheric particulate matter, sulfur dioxide, nitrogen oxides, carbon monoxide, hydrocarbons, photochemical oxidants and lead and established national secondary standards for only particulate matter and sulfur dioxide i2-5). In 1979, EPA set new criteria for pollutants occurring in Laboratory of Microbial Ecology, Department of Biology, New York University, 952 Brown Building, New York, NY 10003. aquatic ecosystems. The Water Quality Criteria for 65 categories of chemicals considered toxic under the 1977 Amendments to the CWA were set at levels considered safe for human health and for various components of the aquatic biota (5-9). These criteria were later defined in terms of 129 specific priority chemicals that are to receive the maximum possible control in the discharge of effluents (10). As there is no "Clean Soil Act", EPA has not formulated criteria or standards for toxicants occurring in terrestrial environments. However, the level of toxicants in soils is indirectly monitored by the Federal Food, Drug, and Cosmetic Act (FFDCA), which requires EPA and the Food and Drug Administration (FDA) to set action levels and tolerances for permissible levels of toxicants in foods. In this manner, some pollutants entering the food chain from con taminated soils and then consumed by human beings are regulated ill). Furthermore, TSCA requires the preproduction testing of any new chemical or any existing chemical with new uses which "may present an unreasonable risk to health or the environment." including both aquatic and terrestrial ecosystems. This article discusses two aspects involved in de veloping standards for toxicants as related to basic environmental safety but not directly to protecting 24# BABICH ASD STOTZKY human health. First, the toxicity of a pollutant to the indigenous biota is dependent, in part, on the physicochemical properties of the recipient environ ment. Because of the large number of chemicals that require testing, it is suggested that microbial assays be utilized as the initial screening system to identify those abiotic factors that influence most the toxicity of different chemical pollutants. Once these interactions have been identified, further testing should be with representative species of the macrobiota. Second, microorganisms in natural habitats, as well as many of the basic ecologic processes that are under the control of microbial activities and which are needed to maintain the quality of the biosphere, are sensitive to pollutants. However, when formulat ing standards for the toxicants mandated by the CAA and criteria for toxicants identified by the CWA, EPA did not consider the potential adverse effects of these toxicants on the various microbemediated ecologic processes le.g., biogeochemical cycles, litter decomposition). Consequently, it is sug gested that the adverse effects of toxicants on mi crobe-mediated ecologic processes be incorporated into the methodologies currently used to set environ mental standards. The development of a new for mulation, termed the EcD* (i.e., the ecological dose 50%, which is the concentration of a toxicant that inhibits a microbe-mediated ecologic process by 50%), would greatly facilitate the incorporation of data on the adverse effects of toxicants to ecologic processes; into the methodologies involved in regula tory legation. Physicochemical Factors: Modifiers of Pollutant Toxicity High Risk Environments Regulatory agencies that establish permissible levels for toxicants in foods, the workplace, and the environment have recognized the existence of hy persensitive subgroups within the general human population. These hypersensitive individuals are termed "high risk groups,** and their sensitivity is determined, depending on the specific toxicant, by such biotic factors as nutritional status, genetic con stitution, developmental stage, and overall health. For example, when setting action levels 'and toler ances for lead (Pbl in foods and milk, FDA recog nized the hypersensitivity of infants and toddlers 12J), and EPA, in its health assessment for cadmium (Cd), stated that "due to increased absorption of Cd being associated with certain nutritional deficien cies, e.g., insufficient levels of dietary iron (Fe), zinc (Zn), or calcium (Cal, older members of the popula tion are likely to be at even greater risk** than the general population tf2). The National Institute for Occupational Safety and Health (NIOSH), in its re view of the scientific literature on occupational ex posure to DDT that was prepared for the Occupa tional Safety and Health Administration (OSHA), noted that "female workers exposed to DDT and other pesticides are reported to have suffered a sig nificantly higher frequency of miscarriages and prepartum disorders than less exposed controls'* US). The toxicity of an environmental contaminant to the biota is influenced, in part, by the physicochemi cal properties (Table 1) of the recipient, environ ment The toxicity of a pollutant may be reduced by the specific abiotic properties of one ecosystem, whereas in another ecosystem with different physi cochemical characteristics, the toxicity of an equiva lent dose of the same pollutant may be potentiated U4-J7). The latter environments should be consid ered high risk environments. High risk groups and high risk environments are essentially similar concepts: a high risk group is a population for which the toxicity of a pollutant is magnified; a high risk environment is an ecosystem in which the toxicity of a contaminant to the indigenous biota is magnified. Consequently, just as regulatory agencies recognize high risk groups when establishing safe levels of con taminants in foods, the environment, and the work place, similar consideration should be directed to identifying high risk environments U8). Table 1. Physicochemical factors of an environment that can affect the toxicity of pollutants,_________ Factor pH (acidity/alkalinity) Eh {oxidation-reduction potential) Aeration status (aerobic, mieroaerobic, anaerobic) Buffering capacity Inorganic anionic composition Inorganic cationic composition Water content Clay mineralogy Hydrous metal oxides Organic matter Cation exchange capacity Anion exchange capacity Temperature Solar radiation Hydrostatic pressure Osmotic pressure Most standards for toxicants are based on a se ries of assumptions, e.g,, that the response of ro dents to a toxicant can be extrapolated to setting a standard for that toxicant suitable to protect human beings; that the response of a few test species to a toxicant can be extrapolated to setting a standard that will protect the multiplicity of life in an entire ecosystem against that toxicant. If the assumptions are either incorrect or incomplete, such as result of the failure to recognize the existence of high risk environ- assump: protect cant th:' mental - tective cal pro; To il factors al pollu ical far metals form, (1 bility. } curs pT wherea as Cdr' cury or wherea curs a* organic fering ria ano marine C X> level o oceans with a OJ -->j -fe. lesser Sev* mobilit mobilii other i carbon for upt of the into sj toxicit; tion of partici phate toxicit; toxic t probat bonate Inor ments of hea on the ent in the he heavy marine ter. Al as Ni2' lated > that c > in its ra tional ex Occupa(OSHA), and "ed a sigand pre- lIS). linant to cochemienviron* lueed by system, it physiequivaentiated considjps and simitar r which igh risk <icity of gnified. cognize > of con3 workzted to ent 5 1 1 ^ t t ' *< i ii a se* of roting a uman i to a idard entire lions jit of risk : { ' I [ STANDARDS FOR ENVIRONMENTAL TOXICANTS 249 environments, then the regulations based on these assumptions will be inappropriate to provide proper protection (19). A standard for an environmental toxi cant that is based on only one set of abiotic environ mental variables may be overprotective or underpro tective for ecosystems with differing physicochemi cal properties (J, 15-17). To illustrate how physicochemical environmental factors influence pollutant toxicity, some heavy met al pollutants will be used as examples. Physicochem ical factors may influence the toxicities of heavy metals by affecting their (a) chemical speciation form, (b) chemical/physical mobility and (c) bioavaila bility. For example, in marine environments, Cd oc curs primarily as a mixture of CdCl*/CdCVCdCl}\ whereas in acidic or neutral fresh waters it occurs as Cd2* (20). Similarly, in marine ecosystems, mer cury occurs primarily as a mixture of HgClf/HgClA whereas in fresh waters, depending on the pH, it oc curs as Hg1*, HgOH", or HgiOHl*. The different in organic speciation forms of these metals exert dif fering toxicities. Fungi tolerated Cd (21) and bacte ria and bacteriophages tolerated Hg (22) better in marine environments or in synthetic media with a level of chlorinity comparable to that occurring in oceans than in fresh waters or in synthetic media with a limited chloride content, thus indicating the lesser toxicities of the metal chloride species. Several abiotic factors limit the chemical/physical mobility of heavy metals. Heavy metals that are im mobilized, e.g., by sorption to clay minerals and other particulates or by precipitation as phosphate, carbonate, or sulfide salts, are less readily available for uptake by the biota. For example, incorporation of the clay minerals, montmorillonite and kaolinite, into synthetic media (23) or soil (24) decreased the toxicity of Cd to bacteria and fungi. The incorpora tion of montmorillonite, attapulgite or kaolinite, of particulate organic matter, or of carbonate or phos phate into a synthetic medium decreased the toxicity of Pb to fungi (25). Nickel and Cd were less toxic to fungi in hard water than in soft water, probably as a result of the higher levels of car bonate and magnesium in the hard water 125, 27). Inorganic cations present in various environ ments may influence the bioavailability and uptake of heavy metals to the biota. Competition for sites on the cell surface between cations normally pres ent in a specific habitat and the cationic forms of the heavy metals may reduce the toxicity of the heavy metals. For example, the toxicity of Ni to marine fungi was reduced in the presence of seawa ter. At the pH and chlorinity of seawater, Ni occurs as Ni*\ and the reduction in Ni toxicity was corre lated with the Mg content of seawater, indicating that competition between Ni and Mg, which have similar ionic radii, for common sites on the cell surface reduced the uptake and, hence, toxicity of Ni (28). The other abiotic factors listed in Table 1 also differentially affect the toxicity of heavy metals (15,16). An environment that may be high risk for one pollutant may be of low risk for a different pollut ant. For example, the toxicity of Cd (24, 29) and Zn (SO) to microorganisms was decreased in acidic sys tems, whereas that of Pb (25) and Ni (31, 52) was in creased. There was no consistent relation between the toxicity of Mn and the pH of the medium (33), and the toxicity of Hg'was pH-independent (SO). The Water Quality Criteria that were suggested by EPA indicate that regulatory agencies have be gun, although only to a limited extent, to recognize that the toxicity of pollutants is dependent on the abiotic characteristics of the recipient environment. In formulating these criteria, EPA noted that "the toxicity of certain compounds may be less in some waters because of differences in acidity, tempera ture, water hardness, and other factors. Conversely, some natural water characteristics may increase the impact of certain pollutants." Consequently, sep arate criteria were set for fresh and marine ecosysterns. Furthermore, as the toxicity of heavy metals appears to be directly related to the degree of hardness in fresh waters, the criteria for Cd, Pb, Ni, Zn, Cu, Cr, and Be were formulated to reflect this "sliding scale", Le., as the hardness increases, tht level erf the metals that can be tolerated by the bio ta also increases (6-5). For example, for hardness lev els of 50,100, and 200 mg/L as CaCO, the criteria for Cd are 0.012, 0.025, and 0.051 pg/L, respectively (9) Although other environmental factors influence the toxicity of heavy metals (as well as of other pol lutants), EPA considered only the level of hardness in fresh waters. For the metals that were evaluated by EPA, the allowable levels were higher in marine than in fresh waters and in hard than in soft waters, indicating that the highest risk, or most fragile, eco systems for heavy metal pollutants would be soft fresh waters. The focus by EPA on only hardness reflects the lack of sufficient data to establish rela tionships between other abiotic factors and pollu tant toxicity. "Although EPA recognizes that other water characteristics such as pH, temperature, or degree of salinity (as in estuaries) may affect the toxicity of some pollutants, the data base at this time is not detailed enough for further specificity". EPA further stated that these criteria will not be "cast in concrete" but will be updated in future years when additional information becomes avail able (6). There is, therefore, a critical need for additional C j2 o ^ ^ 2M> BABICH AND STOTZKY information on the influence of physicochemical fac tors on pollutant toxicity. The continued lack of such data will result in criteria that are inappropri ate (e.g., they will be either under- or overprotectivel. Although the number of abiotic factors (Table 11 and their interactions that can modify pollutant toxicity may appear to be too complex to incorpo rate successfully into standards that can easily be formulated and interpreted, not all these factors are of equal importance in each ecosystem, and not all of the abiotic factors influence significantly the tox icity of each pollutant Most ecosystems possess dis tinct abiotic factors that dominate and serve to characterize those environments. For example, alka line pH and high inorganic ion content are the domi nant characteristics of surface marine waters, and high cation exchange capacity, high organic matter content and acidic pH are the dominant characteris tics of peat soils. Consequently, only the modifying influence of the dominant abiotic factors of specific environments on pollutant toxicity probably need be considered in the regulatory decision-making process. Furthermore, for most chemicals, perhaps only two or three abiotic factors will significantly modify their toxicity. For example, pH and buffer ing capacity appear to be the abiotic factors that most influence the harmful effects of acid precipita tion (74). Consequently, once the dominant abiotic factors that influence the toxicity of a specific pollut ant and the relative importance of these factors in different ecosystems have been established, the en vironmental analyst need focus only on those abiotic factors, ^he establishment of a positive correlation betweerfthe dominant abiotic factors of ecosystems with those abiotic factors that most significantly modify the toxicity of a pollutant (as determined in laboratory screening) should aid in formulating cri teria that would protect all ecosystems against that pollutant For example, if the toxicity of a pollutant is reduced by high pH and salinity, then distinct cri teria should be set for marine and fresh water eco systems, with the latter being the high risk environ ment Conversely, if the toxicity of a chemical is not affected by pH or salinity, one criterion for both fresh and marine water would perhaps provide suit able protection fen1 both ecosystems. Microbes as Assay Systems Most research on chemical toxicants has focused on identifying the effects on human health of both acute and, to a lesser extent chronic exposures and on identifying the molecular bases of the adverse responses. There has been only limited research to evaluate the interactions between pollutants and abiotic environmental factors and the resultant ef fects of these interactions on the general biota. It is the lack of such data that has hindered EPA in set ting criteria that are reflective of the different types of ecosystems in the United States. The volume of chemicals that need such evaluations--e.g., 129 just for the Water Quality Criteria and an estimated 63,000 already in commerce plus approximately 1,000 new ones estimated annually (55), the limitations in laboratory facilities (especially if microcosms are used) and trained personnel, the expensive costs, and the need for Mrapid" results has prompted our recommendation for using microbes as assay systems to identify those abiotic factors that most significantly influence the toxicity of specific chemicals. Microbes can serve as adequate monitors to pre dict the response of the microbiota to a toxicant as influenced by abiotic factors. For example, a compi lation of data of the responses to Cd by representa tives of the aquatic macrobiota (Table 2), terrestrial macrobiota (Table 3) and microbiota (Table 4) indi cates common biologic responses (as well as similar contradictions in data) among these three distinct groups to Cd toxicity as influenced by abiotic fac tors. Microbial assays should be used initially to identify which environmental variables, singly or in various combinations, most directly affect the toxici ty of a specific chemical. Once these variables have been clearly identified for specific chemicals, further studies with representative species of the macrobio ta can be performed, and criteria or standards can be formulated on the basis of their results. The use of microbial assays to predict the toxicity of chemicals to the macrobiota, including human be ings, is not novel. Chronic effects, such as genetic diseases, birth defects and cancer, appear years or even decades after the initial exposure to the toxi cant, and long-term studies using animals must be conducted to detect these latent responses. Such studies are expensive: for example, a single test to determine the potential carcinogenicity of a chemi cal may require as long as three years at a cost of $250,000 or more. Furthermore, the "world laborato ry capacity" for such chronic studies is estimated at 500 chemicals/year, which is not sufficient to keep pace with the 700 to 1000 chemicals introduced an nually into commerce (93). In response to these diffi culties, short-term tests [the best known being the Ames' test (S3)] with bacteria, yeasts, filamentous fungi, plants, insects, and isolated mammalian cells have been developed and are used as rapid and rela tively inexpensive predictors of a chemical's poten tial to cause adverse chronic effects (59). There is, therefore, a need for microbial assays, not only to-screen chemicals for their potential chronic effects on human beings, but also to identify which abiotic factors most influence their toxicity in Tetnperalur- Salinity Water hare Inorganic cations Inorganu _ anions 0rD* <C=J> Organic matiet --J Syntheti CD chela' natura bial ai sions i then ii for pe anima mine shoul< cies c syste Pro Ini Mici At spon right num the > (e-g- EPA in set. "erent types J volume of 129 just estimated lately 1,000 nitations in cosms are 5 costs, and npted our as assay that most f specific rs to preJxicant as - a compi- presenta- errestrial le 4) indi- is similar > distinct iotic faeitiaily to gly or in le toxiciles have . further lacrobiotrds can i i i f i i toxicity man begenetic ears or ie toxilust be Such test to chemicost of x>rato ted at 1 keep ed ani diffig the ntous cells relaoten says, ntial ntify :y in * f r i ii (t> STANDARDS FOR ENVIRONMENTAL TOXICANTS 251 Table 2. Physicochemical factor* affecting the toxicity of cadmium to the aquatic biota._____________________________________ Environmental factorComments_______________________________________________________ ____________________ Reference Temperature The estuarine fish. Fundafua heteroclitus was more sensitive to Cd at 20eC than at 5CC Finger-lings of the freshwater perch,Perea fluviatUit, accumulated more Cd at 15C than at (55) iS7) 5C The estuarine crab. Paragrapsus gaimardii. was more sensitive to Cd at 1SC than at 5C The American oyster. Crassostrea tnrytnica, accumulated more Cd at 20C than at 5C (591 1591 Salinity Increasing the salinity decreased the toxicity of Cd to the grass shrimp, Palaemonetet pugio The toxicity of Cd to the blue crab. CaliinecUt sapidut, decreased with increasing salinity The toxicity of Cd to marine and estuarine crustaceans increased as the salinity was decreased The marine mussel.Afyttfus edulis, accumulated more Cd at 11 than at 30 % salinity Fundulus heteroclitus was more sensitive to Cd at 5 V salinity than ct 15 to 35 M salinity 40! M) U2) 45/ Water hardness The rainbow trout. Salmo gairdneri. tolerated more Cd as the water hardness was increased The fathead minnow, PimephaUs promelas, tolerated Cd better in hard than in soft water The toxicity of Cd to the brook trout, Salvelinut fontinalis, decreased as the water hardness was increased The freshwater snail, Ampullaria paludosa, the catfish, Corydortu punctatus, and the guppy, Lebutet resticulatus, accumulated more Cd in soft than in hard water Increasing the water hardness decreased the toxicity of Cd to eggs of the teleost. Orytias lalipes 44). 45) 46/ 47) 481 49) Inorganic cations Simultaneous exposures to Pb, Zn, and Cu reduced the uptake of Cd by the freshwater plant, Elodea nuttallii Ca decreased the toxicity of Cd to the marine amphipod, A/crinopamtnaruf obtuxatus Zn reduced the toxicity of Cd to Pimephales promelas Ca reduced the toxicity and uptake of the Cd by the freshwater shrimp, Ganfflanu pulex (5(7) ISJ\ iSS (55) UAL 03677 Inorganic anions Pyrophosphate reduced the uptake of Cd by Daphnia magna (54) Organic matter Colloidal orgaoic particulates decreased the toxicity of Cd to the freshwater crustacean, Stmocephalus lemtlatus Humic acid reduced the uptake of Cd by Croxxoxtrea trirptntca and by Daphnia magna (55) ISS, 54) Synthetic chelators NTA reduced the toxicity of Cd to Palaemonetet pugio EDTA and NTA reduced the uptake of Cd by Daphnia magna and by Crastoetrea virginica EDTA. NTA. and DTPA reduced the uptake of Cd by the carp, Cyprinut carpio EDTA reduced the uptake of Cd by the marine barnacle, SemtbaiaRux balanoidei 4(7) (59.54) 156) (57) natural environments. Just as the results of micro bial assays are used to make more informed deci sions as to which chemicals should be examined fur ther in the limited number of laboratories equipped for performing chronic toxicity studies with whole animals, microbial assays should be used to deter mine which abiotic factor-pollutant interactions should be studied further with representative spe cies of the macrobiota in either simplified artificial systems or in complex microcosms. Protecting the Environment In Toto Microbe-Mediated Ecologic Processes Attention by environmental policy-makers re sponsible for regulating toxicants has focused, and rightfully so, on human health, as evidenced by the numerous federal statutes concerned with limiting the exposure of human beings to harmful chemicals (e.g., CAA, CWA. FIFRA, FWPCA, FFDCA, RCRA, TSCA). However, the continued health and welfare of human beings is dependent on maintain ing the quality of the biosphere, as acknowledged in TSCA, which requires the preproduction testing of new chemicals and the testing of existing chemicals with new uses for their potential hazards to the en vironment As stated in TSCA, MIt is the policy of the U.S. that adequate data should be developed with respect to the effect of chemical substances and mixtures on health and the environment." Reg ulatory agencies and environmental policy analysts appear to have narrowly defined ``effect on the environment" as direct effects on the biotic compo nents of the biosphere and have not considered the effects of pollutants on ecologic processes mediated by the biotic component and which are necessary to maintain the present state of the environment. For example, EPA has stated that the Water Quality Criteria were intended "to reflect the latest scientif ic knowledge on the identifiable effects of pollut ants on public health and welfare, aquatic life, and 252 Environmental factor pH Temperature Salinity Cation exchange capacity Water content Nitrogen content Inorganic cations Inorganic anions BAB1CH AND STOTZK Y Table 3. Physicochemical {actors affecting the toxicity of cadmium to terrestrial plants. Comments___________________________________ ______________ _____________ -__________ Reference Uptake of Cd by oaU and lettuce increased as the pH was decreased Uptake of Cd by corn was independent of soil pH 156' _ <5P' Increasing the soil pH from 5.5 to 7.5 reduced uptake of Cd by rice Chard and tomato accumulated more Cd when grown in acidic (pH 5.0 to 5.7) than in alkaline (pH 7.5 to 7.8) soils Increasing the soil pH from 4.5 to 6.4 reduced the uptake of Cd by ryegrass and oat 16Q> ($2' <62' Uptake of Cd by soybeans increased as the soil temperature was increased (66i Increasing the salinity from 0 to 10 < decreased, but from 10 to 30 increased, the toxicity of Cd to germination of seeds ofSpartina altemiflora f6i> Uptake of Cd by ost was lower in soils with high than with low cation exchange capacities Increasing the water content of the soil increased the uptake of Cd by barley No synergistic interaction was noted between a drought stress and Cd for growth of An- dropogon aeoparius. Monarda fistulota, and Rudbeckia kirta Uptake of Cd by fescue, grown in soil, was enhanced by nitrogen amendments Uptake of Cd by bush bean, grown in a nutrient solution, was decreased by nitrogen amendments Uptake of Cd by oat and lettuce, grown in a nutrient solution, was decreased by the addition of Ca. K. or A1 Synergism was noted between Cd and Pb in reducing root growth, woody stem diameter growth, and foliage growth of American sycamore Synergism was noted between Cd and Pb in reducing vegetative growth of corn shoots A1 reduced the uptake of Cd by Holms lonofus Ni or Pb added to soil increased the uptake of Cd by ryegrass Uptake of Cd by oat, grown in soil, was decreased by the addition of phosphate Phosphate amendments decreased the uptake of Cd by corn seedlings - J65) (66) (87) (68) (68) (56' 176) (72) (721 1621 (76) (741 Table 4. Physicochemical factors affecting the toxicity of eadmiom to the microbiota. Environmental factorComments pH Increasing the pH from 5 to 9 progressively increased the toxicity of Cd to Aspergillus m'per, from pH 7 to 9 increased the toxicity of Cd to Bacillus cere**, Aicaligenes foecolis, and Trichoderma viride, and from pH 8 to 9 increased the toxicity of Cd to Agrobacterium tume/adens, Nocardia panffinae. ndRhitopus tolon%fer\ pH did not affect the toxicity of Cd to Streptomyces oltvaeeta Increasing the soil pH from 5.1 to 12 increased the toxicity of Cd to mycelial growth of Aspergillus niger but not of Aspergillus fischeri Increasing the pH from 6 to 8 increased the toxicity of Cd U> Micrococcus luteus, Stapkylococcus aureus, Clot fridium perfringeru, Escherichia coli and Pseudomonas aeruginosa; pH did not affect the toxicity of Cd to Bacillus rubtilts Increasing the pH from 6.5 to 8.3 increased the toxicity of Cd to Chlorella pyrenoidosa Cd toxicity to Chlorella pyrenoidosa decreased as the pH was increased from 7 to 8 Uptake of Cd by the diatom, Navicula pyrenoidosa, and the green alga. Chlorella pyrenoidosa increased as the pH was increased from 6 to 8 Increasing the pH from 6 to 9 decreased the toxicity of Cd to the cyanobacterium, Nostoc calcicola The pH-Cd toxicity interaction towards mycelial growth of the fungi. Achyla sp. and Saprolegnia sp., was dependent on the composition of the growth medium Temperature Chlorella pyrenoidosa accumulated Cd faster at 25C than at 4C Chlorella pyrenoidosa accumulated Cd faster at 15C than at 5C Water hardness The alga, Sitellaflexilis, accumulated more Cd in soft than in hard water Rhizopus ttolonifer, Scopulariopsis brevicaulis, PeniciUium uermieulafum, Trickoderma viride. JJeauvena sp., and Aspergillus niger tolerated Cd better in hard than in soft water Reference (28) (24) (75) (76) (77) 178) (781 (661 176/ ISO) (46) (26) Environments factor___ Salinity Synthetic chelators Organic matter Clay mineral- Cation ex capacil C 3D r*" Inorgani' O cation* CO --I CD Inorgar anion recre limit* this vers* \y or forir fiabl micr M tem ic p (62) (65J (64/ (65/ (66) (67) (66) (691 (56I (701 (7/J (72) (621 (73) (74) srence 29) 4) 5) 7) n ) Environmental factor Salinity Synthetic chelators Organic matter Clay minerals Cation exchange capacity Inorganic cations Inorganic anions STANDARDS FOR ENVIRONMENTAL TOXICANTS Table 4 (Continued) 253 Comments Increasing the salinity above 45 /oo reduced the toxicity of Cd to an unidentified marine bacterium The toxicity of Cd to Rhizoptu stolonifer, Trichoderma viride, Aspergillus niger, and Artkrobotrys conoides was reduced in medium amended with seawater at 20% or greater EDTA decreased the toxicity of Cd to the marine diatom, TXtyfam brighttoellii NTA reduced the toxicity of Cd to photosynthesis of a natural freshwater phytoplankton community EDTA reduced the toxicity of Cd to Klebsiella pneumoniae EDTA reduced the toxicity of Cd to Nostoc calcicola Pyruvate, gluconate, citrate, and aspartate reduced the toxicity of Cd to Klebsiella aerogenes Increasing the concentration of peptone decreased the toxicity of Cd to an unidentified marine bacterium Citrate increased the toxicity of Cd to Pseudomonas sp. but not to Etekerickia colt Glutamine and cysteine decreased, but citrate increased, the toxicity of Cd to Nostoc calcicola Humus reduced the toxicity of Cd toSeianastmm capricomutum Montmorillonite and. to a lesser extent, kaolinite decreased the toxicity of Cd to BactUus megaterium, Agrobacterium tumefaciens, Nocardia coralline, Fomes onnotus, Pkoliota margtnata, Botrytis cinerea, Aspergillus niger, Phycomyces blaketleeanut, Trichoderma viride, Chaetomium sp., Thielaviopsis paradoxa, Scopularioptis brevicaults, and Sehisophyllum sp. in synthetic medium Montmorillonite and. to a lesser extent, kaolinite protected PeniciUtum vermiculatum, Aspergillus asperum, Aspergillus niger, Aspergillus fisekeri, and Trichoderma viride against Cd toxicity in soil Cd was less toxic to PeuiciWim vermiculatum, Penicillium asperum, Aspergillus niger, Aspergillus fisekeri, and CuwninghameUa eckinulata when grown in an alkaline soil with a high cation exchange capacity (j.e., 16 meq/100 g) than in an acid soil with a low eation exchange capacity (i.e.. 8.2 meq/100 g) Mg reduced the toxicity of Cd to growth of Escherichia coli Se reduced the toxicity of Cd to growth of Haematococcus capensis The toxicity of Cd to growth ofAspergillus niger was decreased by Ca and Mg Zn decreased the toxicity of Cd to growth of Euglena gracilis Mn inhibited the uptake of Cd by CkloreUa pyrenoidota Cd and Pb interacted synergistieally towards inhibiting growth of a brackish water phytoplankton community Cd and Pb interacted synergistieally to inhibit photosynthesis and nitrogenase activity in Anabaena tnequalis Zn and Pb interacted synergistieally, but Hg and Ni interacted antagonistically, to Cd- induced mitotic delay in Physamm polycephalum Zn and Cd interacted synergistieally to inhibit growth of the marine diatoms. Thaiassiosira pseudonana and Skeletonema tricomutum; Zn interacted antagonistically to the toxicity of Cd to growth of Skeletonema costatum Cd'* was more inhibitory than was an equivalent concentration of Cd as CdlCN),'* towards growth of a mixed microbiota from activated sludge Increasing the chlorinity decreased the uptake of Cd by the estuarine alga, CkloreUa salina Chloride, at a level equivalent to that occurring in seawater, decreased the toxicity of Cd to mycelial growth of Sepedontum sp., Oospora sp., Trichoderma viride, Aspergillus niger, Rhizoput stolonifer, and Scopularioptis brevicaults__________________________________ Reference (61) (SI) <821 (65) (64) (791 (641 (SJ) (65) (79) (66) (S3) (24) (SO) (67) (661 (69) (90, 91) (60) (921 (951 (94) (95) (96) (97) (SI) recreation'' When considering "aquatic life," EPA limited the scope to animals and plants, including in this category the unicellular algae (}. As the ad verse effects of toxicants on the microbiota, primari ly on bacteria and fungi, were not considered when formulating these criteria, EPA ignored the "identi fiable effects" of these toxicants on the numerous microbe-mediated ecologic-processes. Microorganisms in aquatic and terrestrial ecosys tems are dynamically involved in many basic ecologic processes, such as the biogeochemical cycling of chemical elements, the mineralization of carbon, ni trogen, sulfuT, and phosphorus needed to maintain the fertility of the biosphere, the formation of or ganic matter by chemo- and photosynthesis, and the decomposition of plant and animal wastes. The hin drance of these microbe-mediated ecologic pro cesses by anthropogenic pollutants would greatly affect the quality of the biosphere eventually adversely affecting human health and welfare. For example, microorganisms, primarily fungi and bac teria, are involved in the decomposition of organic URL 03619 2.'>1 BABICH AND STOTZKY matter, such as complex animal and plant tissues and excretory products. In addition to being "Na ture's sanitary engineers,'1 microbial conversion of organic matter to inorganic materials (i.e., mineral ization) is an important nutrient regeneration pro cess in aquatic (100) and terrestrial (101) ecosystems. Although most natural ecosystems contain an abun dant supply of carbon, nitrogen, sulfur and phospho rus, the major portion of these elements occurs as organic complexes that, as such, are unavailable for uptake by the phytobiota (102). Reductions in the mineralization activities of microbes would initially affect the primary producer level, with plant growth being limited. As plants are the basic com ponents of all food chains and webs, such perturba tions in plant growth would hinder the population dynamics of herbivores, carnivores and omnivores, including human beings. Thus, an adverse effect on a microbe-mediated ecologic process such as miner alization would, by a "domino effect,'* eventually im pinge on the continued health and welfare of human beings. Microbes are sensitive to most pollutants (lb-17), and an inhibition of microbial activity is accom panied by reductions in the ecologic processes that they perform. The adverse effects of toxicants on microbe-mediated ecologic processes have not, as yet, been incorporated into the formulations for computing criteria and standards of environmental risks If5, 103). For example, although Cd adversely affects many microbe-mediated ecologic processes (Table 5), EPA did not consider these processes when formulating the Water Quality Criteria for this metal (4) or for other toxicants (6-8). The need to examine environments in a "holistic framework," including microbe-mediated ecologic processes, has been noted as a goal in the 1980s for environmental analysts (105). It is difficult to understand the failure of environ ments] policy analysts and policy makers to con sider the adverse effects of toxicants on microbemediated ecologic processes when formulating crite ria such as the Water Quality Criteria and stan dards such as the National Secondary Air Quality Standards. The failure may be due to the inability to compare easily, and, thus, to evaluate and incorpo rate into the existing methodologies used to com pute environmental criteria and standards the ex tent of damage by a toxicant to an ecologic .process in different types of ecosystems. More probably, en vironmental toxicology has simply not developed to the point where the need to consider an adverse af fect on an ecologic process is appreciated. It has been stated that aquatic toxicologists have only be gun to address the "ecological effect" of toxicants um Ecologic Dose Fifty Percent (EcDM) The extent of pollutant damage to some microbemediated ecologic processes can be measured effec tively in the laboratory. For example, heavy metals have been shown to interfere with several microbemediated ecologic processes, such as the biogeo ehemical cycling of nitrogen (115, 122-135), sulfur (107), phosphorus (108, 133, 134), and carbon (108, 109, 111-115, 129, 136-138Y, the decomposition of plant litter (SO, 109,110,117-119,133,139k photosyn thesis (S3, 92, 115,121,140k and enzymatic activities (11, 119, 131-134, 141, 142). As these adverse effects on ecologic processes can be quantified, it b sug gested that a formulation be derived, similar to the LD* (i.e., the dose that is lethal to 50% of the ex posed population) which has been used extensively to compute standards for exposures of human be ings and the general biota to toxicants (143), to al low environmental analysts and policy-makers easi ly to compute the extent of damage by a toxicant to a microbe-mediated ecologic process and to compare the extent of damage by the same toxicant to a common ecologic process in different types of eco systems. Such a formulation, termed the "ecologic dose fifty percent" (EcD*) and defined as the dose of a toxicant that decreases a specific microbe-medi ated ecologic process by 50% (other percentages of decrease could also be used), would permit regulato ry agencies to incorporate such data into the exist ing methodologies used in establishing environmen tal criteria and standards (18,103). The EcD* can be determined in a manner similar to that used for the LD*, in which a population, or in the case of the EcD*, a microbe-mediated ecologic process, is exposed to progressively increasing lev els of a toxicant. The resulting data, when plotted as percent mortality for the LD* or as percent inhi bition for the EcD* versus the concentration of toxi cant, should approximate a broad $*shaped curve from which the LD* (144) or the EcD* can be com puted. The LD* test, which was developed initially in 1927 for the biological standardization of hazard ous drugs, has been incorporated into the routine toxicological protocol for other classes of chemicals and now is part of practically all Federal guidelines that regulate the toxicological testing of chemicals (1451 Currently, toxicologists determine LD* values of environmental chemicals for plant and animal species representative of specific ecosystems, and then, environmental policy-makers utilize the LD* values of the mo6t sensitive species as the bases on which to formulate criteria. Similarly, EcD* values could be computed for different ecologic processes stressed by a common pollutant, and the EcD* value of the most sensitive microbe-mediated eco- Ta*1 Ecologic process Soil enzymai'.' activity Carbon mineralize: Litter decompos C 3rD* Microbial o photosyr CO CO COO Nitrogen c Denitn Nitrifi' Nitrogen fixat i< logic pi 103). The LD*- F species iologic; fore, c popula specie; protec media' combi) ria an combi ^ ... - t asSbta URL 03681 STANDARDS FOR ENVIRONMENTAL TOXICANTS 255 Table 5. Effects of cadmium on aome microbe-mediated ecologic proceaaea in aquatic and terrestrial ecosystems. nicrobe*d effec- ' metals nicrobebio^eo- sulfur >n U08, tion of rtosyn- ivities effects s sugto the he exsively tn beto aJ5 easiant to npare to a f ecoilogic e of medies of ilatoxistmen- ' ! f Ecologic process Soil enzymatic activity Carbon mineralization Litter decomposition Microbial photosynthesis Nitrogen cycle Denitrification CommentsReference 25 nmole Cd/g soil inhibited arylsulfatase activity 25 pmole Cd/g soil inhibited the activities of acid and alkaline phosphatases U07) 008) JO ppm Cd inhibited soil respiration Soil respiration was decreased by addition of 10 ppm Cd + 1000 ppm Zn Starch decomposition and soil respiration were reduced in a spruce needle mor con- laminated with Cu. Zn, Pb, and Cd emitted from a brass foundry Carbon mineralization in soil was inhibited by 100 ppm Cd 1000 ppm Cd extended the lag phase of glucose degradation in soil; no synergistic in- teraction was noted between 1000 ppm Cd and up to 10,000 ppm Zn or simulated acid rain causing a reduction in soil pH to 2.6 or 3.2 to glucose degradation-* 10 ppm Cd inhibited glucose oxidation in Chesapeake Bay water and sediment - 009) (110) (111) (112) (113, lliI (115) Rates of decomposition of spruce needle litter obtained from sites near metal-processing industries emitting Cu, Zn, Ni, and Cd were reduced as compared to litter obtained from nonpolluted sites Decomposition rates of leaf litter from Guerras ueiutina Smilacinc ttellata. and Populus tremuLndei were reduced in a site contaminated with Cd. Zn, Pb and Cu Decomposition rates of litter consisting of leaves from Sattafrai albidum, Guerras prinut, and Guerras rubra and contaminated with Cd, Cu, Fe, Pb and Zn were lower as com pared to similar Utter from a nonpolluted site 1000 pg Cd/g soil inhibited decomposition of a Douglas fir needle litter Decomposition of leaves of Pinus taeda. Sastafras albidum, Guerras nigra, Quereus iauri/olia. Prunut americana, and A eer rubrum was decreased in a freshwater ecosystem amended with 5 mE Cd/L (116) (117) (118) UJS) (ISO) 0.1 mg Cd/L reduced photosynthesis of a brackish wster phytoplankton community 25 ppm Cd inhibited photosynthesis in Chesapeake Bay water 1001iM Cd inhibited growth of a marine phytoplankton community 10-J/ Cd inhibited photosynthesis of a freshwater phytoplankton community consisting mainly of diatoms (92) 015) (121) (&?) Denitrification by the indigenous microbiota was reduced by 100 Mg Cd/g soil (122/ nilar l, or ogic levtted nhioxirve om- illy .rdine als les als tes lal nd 3. an ss 3S Vi O- >% "* Nitrification Nitrogen fixation 0.01 to 0.04M Cd inhibited nitrification in soil Nitrification w*$ reduced by Cd concentrations up to 400 mS Cd/g soil but was enhanced at levels from 400 to 2.500 Mg Cd/g soil 5 nmole Cd/g soil inhibited nitrification 500 ppm Cd reduced nitrification in soil; at 1000 ppm Cd. nitrite accumulation was evident Nitrification was reduced in Chesapeake Bay water amended with 100 ppm Cd l&fiM Cd inhibited nitrogen fixation by soybean nodules containingftfciiobtum japonicum Nitrogen fixation ofa Dougtas-fir needle litter was decreased by amendments of 5mM Cd/g soil (123/ (12k) (1251 (126) (US) (127) (85) logic process could be used to formulate criteria U8, 103). The EcDjd has three distinct advantages over the LD. First, LD values are for populations of single species, which usually are of uniform size, age, phys iological and genetic constitution, eta, and, there fore, do not display the heterogeneity of natural populations IM). Standards based on such single species populations may not, therefore, adequately protect the biosphere. Conversely, most microbemediated ecologic processes are controlled by the combined metabolism of different species of bacte ria and fungi, and thus, an EcD* value reflects the combined response of a variety of populations to a stress. Second, the species selected to be assayed in LEV tests may be of limited importance to a natural ecosystem, and when toxic effects are noted, deter minations must then be made as to whether the presence of the species is critical to the continued functioning of the ecosystem. However, a greater risk and perturbation to the functioning of an eco system would be the inhibition or removal of an en tire functional group, such as decomposers, nitrogen fixers, or primary producers 1147). The determina tion of EcD values would, therefore, have more rel evance than would LD* values for predicting the continued functioning of stressed ecosystems. Third, with the LD* test, a direct comparison be- 256 BABICH AND STOTZKY tween the sensitivities to a toxicant of species that dwell in different ecosystems is not always possible. For example, it may be necessary to compare the sensitivity to a pollutant of a marine and a fresh water fish. These comparisons are difficult, as the possible effects resulting from the differences in the environments are confounded by differences in the test species. However, as most microbe-mediated ecologic processes are common to all ecosystems, a reduction in a process in one ecosystem by a toxi cant can easily be compared with a similar reduc tion by that toxicant in the same ecologic process but in a different ecosystem. For example, the level of toxicant inducing a 50% reduction in carbon min eralization in fresh waters can be compared to the level of that toxicant evoking an equivalent reduc tion in carbon mineralization in marine waters CI& 103). Although it is suggested that the EcDw concept be incorporated into regulatory decision-making, it is recognized that this concept needs to be more ful ly analyzed and developed by the scientific commu nity. For example, a 50% reduction in a basic ecolo gic process may be a value that is too extreme for the continued functioning of a perturbed ecosystem and, perhaps, an EcD* or EcDw would be more suit able. Also, the EcDs of a specific ecologic processpollutant interaction should not be viewed as a con stant value, as the EcD* value may depend on the lengthjof exposure and on the properties of the test ecosysfem. For example, an EcD value determined after 2 days of exposure, during which a temporary lag may occur in the ecologic process being studied, may be entirely different if determined after 2 weeks of exposure, during which time the stressed populations may have adapted to the toxicant or may have been replaced by populations having com parable metabolic capabilities (113,114,126,148). An EcDg, value for an ecologic process-pollutant interac tion may be different for hard fresh waters than for soft fresh waters. These "problems" are not unique to the EcDk but also apply to the LD*, and it is com mon for toxicologists to determine an LDb or LDM or to determine an LD after 24, 48 or 96 hr or even after 2 weeks of exposure. Although not often em phasized, the LD is also not a constant hut is depen dent on or, at least, influenced by species, age, weight, sex, genetic constitution, health, diet, meth od of exposure, ambient temperature, seasonal vari ation, etc. (145). Another aspect that will require considerable de velopment is the application of appropriate statisti cal designs and analyses to the ecologic data used for calculating EcD values. This problem is also not unique to the EcD concept, as the appropriate sta tistics for LDb data and risk levels of carcinogens and other environmental chemicals are still being debated (19, 149-158). The accumulation of sufficient data and numerous attempts to apply the EcD concept should, with the aid of statisticians, resolve this problem. The EcDa> concept can be applied to many areas of environmental toxicology and is not limited to the Water Quality Criteria. There have been few legislative or regulatory initiatives designed to pro tect soil as an ecosystem, even* though pollutants may cause serious adverse effects on microbe-medi ated ecologic processes in terrestrial ecosystems. Consequently, the implementation of EcDu values in risk analysis of aquatic ecosystems should have immediate application to terrestrial ecosystems sim ilarly stressed by pollutants and, thus, may result in the establishment of Soil Quality Criteria. Conclusions Cairns (146), in discussing future needs in the bio logic assessment of pollutants, mentions two con cepts: "pollutant realism" and "environmental real ism." Pollutant realism is attained when those char acteristics of the test compound that exist in the natural environment are incorporated into the labo ratory test system. As EPA has begun to recognize that the physicochemical properties of the recipient environment influence the toxicity of a pollutant to the indigenous biota, such abiotic factors should be routinely considered when formulating environmen tal criteria and standards. However, at present, the data base for such interactions is insufficient, and laboratory tests using animals and plants are too tedious and expensive. As the influence of abiotic factors on the response of microbes to pollutants is similar to that exhibited by more complex systems (i.e,, plants and animals), it is suggested that micro bial assays be used initially to identify those abiotic factors that most influence the toxicity of the vari ous pollutants. Once these factors have been de fined, additional studies should be performed with these factors using macrobiotic species representa tive of the stressed ecosystems and then criteria and standards formulated. Environmental realism is attained when the tests account for all aspects of the ecosystem, including those ecologic processes controlled by microbial activities. Microbe-mediated ecologic processes are critical to the continued func tioning of the biosphere, and some of the environ mentally oriented Federal statutes, such as TSCA, specify that adverse effects of pollutants on the en vironment must be determined. Thus, it is also rec ommended that these ecologic events be considered in the regulatory process, and it is further sug- being "iicient EcDk esolve areas led to n few o proitants meditems. ies in have simjlt in bioconreaJrharthe abonize ient t to be lenthe md too >tic > is ms rotic Tiieth aia is of iS d C> t, 1- i STANDARDS FOR ENVIRONMENTAL TOXICANTS 257 gested that an EcD* formulation would be a useful tool to simplify their incorporation. 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