c: n"-~ .F-~'h F" "T E F: ][ I

LITERATURE REVIEW

The toxicity of he~vy metals in aquatic system is influenced

by many environmental factors like pH"of the medium, pE, presence j

of other inorganic and organic chemical species, temperature et.:.

Mo~t of these factors effect the toxicity by their influence on

the speciation of the pollutants in the system (Babich and

Stotzky, 1983a; Borgmann, 1983). Klotz (1981) states that these

factors include those which influence the biological availability

of the metals and those influencing the physiological conditions

bf the organism. The toxicity of heavy metals depends upon the

chemical fractionation of these element~. A brief review of the

literature on speciation is very important before going to

toxicity.

Spec i at i on connots to the di f ferent physi co-chen"li cal forms

of an element which make up the total concentration of an element

in a system (Florence, 1982; Cross et ala 1984). A number of

studies have been reported that' mainly: the to)<;ic form of the

metal is the bio available form, and it is reported by many

workers that the toxic form rather than the total metal present

should be considered in all toxicity studies (e.g., Pagenkopf ~~

~l. 1974; Andrew ~i ~l. 1976; Sunda and Guillard, 1976; Anderson

and MC1r_el, 1978 Bunda ~i ~l. 1'378; Bunda and Lewis, 1978; Gachter

1978; Sunda and Gillepsie, 1'379; All en ~:t. 21·, 1 '380;

Huntsman and Sunda, 1980; Peterson, 1982; Geisy ~:t. 21. 1983;

Laoma, 1983) . Nurnberg (1983) objects to the general assumption

among ecotoxicologist of giving importance to only dissolved

metals and labile complexes CMeXj) including fre~ hydrated heavy

cations (Me"·) and argues that ~t any rate the strong metal

heavy metal complexes (MeLm) favours the adsorption and

consequent uptake of heavy metals by organisms. However, it has

been reported from thermodynamical considerati6ns that the metal

protein interactions which lead to the transport of metals across

the membrane,

or Cu(OH)+ in case of bivalent (e.g., Chakou~akos ~:t. ~l· 197'3 -

the difference in toxicity of.different form of Cu) ions and for

trivalent ions like Fe iii, organically bound form - as these

cations get inactive due .to hydrolysis and polymerisation - will

be favoured (Cooley and Martin, 1980).

Speciation of heavy metals in natural systems has been

studied mainly based on two different techniques (a) chemical

modelling and (b) ex~erimental methods (e.g., Stiff, 1971 ;

Gardiner, 1974; Sylva, 1976; Vuceta and Morgan, 1978; Jackson and

Morgan, 1978; Wilson, 1978; Mouvet ~t ~l. 1982; Mouvet and Bourg,

1983; Nurnberg, 1983, 1984; Park and Allen, 1984) • Chemical

modelling studies (computer) involved the use of known st~bility

constants along with con':entration of yarious ions and suspended

12

solids in the water~ The main obstacle to this technique is the

lack of reliable thermodytlarJ"li c data (FI or enc e, 1982).

Experimental techniques engaged in, are ion selective electrodes

(Stiff, 1'j71 ; Gardiner, 1'374; Shephard !t!. ~!.. 1'380; McGrath ~!

1984) and anode stripping voltammetry (Brezonik, 1974; del

Castilho ~t. ~l· 1983; Nurnberg and Valenta, 1'383; Nurnberg,

1983,1984; Sugawara ~i ~l. 1984; Valenta ~i ~l. 1984).

Upon the entry of a metal into an aquatic system it is

partitioned into many forms depending upon the characteristics of

the system like pH, redox conditions, solubility,

concentration of other metals and compl~xes,

presence and

H1e elements

are partitioned into liquid and solid phases and in each phase

itself further partitioning occurs in between specific legends

present in the medium and the process is determined by the

concentrations of the legand and the strength of each metal

legand association (Vuceta and Morgan, 1'378) • The speciation is

a net result ~f the interaction of all constituents of an aquatic

system through complex formation and s,::,lid precipitation

1973), involving the solute solvent interactions,

hydrolytic reactions of the elements, complexation reactions and

the reactions involving solid and solutio~ interfaces (Lac~ie and

James, 1974).

It1 natural -aquatic systems Singh and Subramanian (1984)

summarise that heavy metals remain in si); phases, vi z. , (1)

13

diSSQlve-d phase- including dissolve-d fre-e- _ i ()ns, dissolve-d

inorganic and organi t: comple-}';e-s; (2) colloidal/ suspe-nde-d phase

including inQrganic and organic metal colloids; (3) sorbed site-s

including clay e-xchange site-s and sites in metallic hydroxide-s;

(4) carbonate-s; (5) detr i tal mi neral s, and (6) crystall ine si t~s.

Stumm and Bilinski (1972) put forward another scheme- of metal

speciation mainly based on the- particle size-. According to this

scheme, it can be distinguished into mainly two par t s, viz.

filte-rable- and non-filterable, the limit of filte~able being 0.45

mu pore size me-mbrane filters. Guy and Chakraborti (1975)

distinguish the dissolved metal ions into (a) si ng 1 e aquat e-d

me-tal ions; (b) me-tal ions comple-xed to organic compounds like

fulvic, humic and amino acids.

Redox conditions influene the trace metals in aquatic system

by way of inducing direct changes in the oxidation state of the

metal ion and also by making t:hanges in available- and compe-ting

ligands or chelater. The sediment water interface in natural

waters is one- of the sites at which inte-nse ~edox activities take­

place- due Inainly to the- deposition and accumulation of organic

matte-r and the difficulty of mole~ular oxygen to diffuse down

into the sediment interstitical waters. pH influences the

speciation significantly. As the hydrogen ion concentration is

decrease-d carbonate; oxide or· even silicate- ions become the

predominant specie-s (Morel et ~l. 1973).

1~

From an eco-to;,;icological point.of view, Babich and Stotzky

. (1983a) state that the hydrogen ion concentration affects th~

toxicity through alterations in bioavailability by its direct

influence on the chemical speciation - of each metal and also by

the alteration of the complexing capacity of different compounds

with the metal. Khalid ~t fi!!... (1978) report about the changes

brought about by oxygenation of sediment in pH and redox

potential and its influences on transformation of heavy rrietals.

They observed strong modification in the distribution of trace

metals in various chemical fractions of the sediment due to the

increase in redox potential and reduction in pH brought in by

o;,;ygen purging. They observed an increase of 25-30% of Pb, Cd

and Cu in water soluble fractions due to treatment. Another

situation noted was the immobilization of metals like Mn

and iron CFe 2 + by their oxidation to Mn 4 + and Fe3 '" forms) by the

increased redox potential. Mantoura ~t ~!... (1978) discuss about

the influence of salinity on speciation of different heavy metals . 1 ike Cu, Cd, Zn, Mg, Mn, etc~ It was noted by them that ionic

metal and metal organic complexes decreases with increase in

salinity while hydroxy, chIaro, and sulfate species increased.

Heavy metals show varying degree of hydrolitic reactions in

aquous medium. All metal cations are hydrated in water and the

coordination reactions in which these cations take part are

exchange reactions of the coordinated water with other 1 igands.

15

Highly.,i(lnic inter,a,cti~ns like),ydrations of metals' are found' .. ";' \, ' .• ;. ~ '. . 1;' ."

to

haveco~relationwith the formal charge 61 the metal ion_CZ) and

its ionic radius Cr), or the hydrolytic parameters like hydration

are found to vary with electrostatic energy ratio CZ2/ r ). A

gradual change from aquohydroxo, hydroxo-oxo and oxo-complexes as

the pred':Jminant species in the pH range. of aquous solution with

formal charge of the cations is generally observed (Stumm and

Morgan, 1'370). Strongly charge~ elements like Co+3 are strongly

hydrolyzed, divalent Cu2 +, Ni~~'hydrolyze in ~he range of natural .... {.

waters (6-12) while alkaline earth elements only hydrolyze in

basic solutions (Lackie and James, 1974) . Mul t i nUt: 1 ear

hydrolysis products are commonly seen in case of metallic

cat i ':Jns. Stumm and Morgan (1970) propose the following rules for

the formulations of hydrolysis reactions equilibria: the

tendency of metal ion solution to protolyze (hydrolyze) increases

with dilution and decresing pH; (2) the fraction of polynuclear

com~lexes in a solution decreased on dilution.

The main inorganic complexing legands to heavy metals in

water are: Cl-, S042-::, HCD3-, ,sulfide and phosphate species

(Lackie and James; 1974). The combination of legands with

cations are highly selective 'and the preference is a function of

cations in aqueous solution. Based on the preferenc~ of legands t.· '

by cation, Ahrland gi 21.· ('1~58) ~l~ssified metal ions into

different groups. Nieboer 'and Richardson (980)'s classificati':Jn

is a further development of Ahrland gi 21. (1'358) ttl v'iew 6f the'

• I

16

physico-chemical characteristics of the ions and their role in

toxicity mechanisms. The metal ions of class A are visualised to

be of spherical symmetry and low polarizability (hard spheres)

while those of class B are of low polarizability (soft spheres).

In cas~ of class A elements they are typical oxygen seeking and

they form no suI fide precipitates or complexes as OH- io-ns are

bound before HS- or S2- groups. Their complex stability are

explainable by simple electrostatic picture of the binding of

cations and legands. On the other hand, class B elements are in

general nitrogen/sulf~r seeking. They form insoluble sulfides

and soluble complexes with S2- and HS-.

The concentration of inc.rganic legands determine to a great

extent the distribution of metals between solids and solution

phases. Many·; of the metal legands complexes formed, li ke

carbonates and sulfides are least soluble in water and get

precipitated leading to a decrease in the net dissolved metals

content. It has been observed in case of sulfide interactions

with metals that as sulfide ions have more affinity to iron, the

formation of iron sulfide alters the release or ~etention of

other metals in anoxic conditions. Some metals are more soluble

in anoxic conditions because of the higher ~olubility of the

respective sulf,ides (Fe, Co, etc.), while metals like Cd, Hg etc.

are less soluble (Singh and S~bramanian, 1984)& But it has been

reported'that by an increase in pH and HS- or S-2 content, metals

like Hg again become soluble.

17

Adsorption to inorganic surface particles like clay r,linerals

is another factor which determines the partitioning of heavy

metals in aqueous syste~ (Bourg, 1'383; Morel §:t ~l· 1'384) • A

s~eciation study ,conducted by Mouvet and Bou~g (1983) on sediment

of Meuse river emphasizes the imp6rtance of adsorption process in

the control of trace metals concentra,tion. The uptake of aqueous

metal ions are generally attributed to a number of processes like

adsorption, ion-exchange and co-precipitation, and cOLllumbic

interactions with the surface of inorganic particles in its

double layer. Clay minerals at the range of natural pH of the

ecosystem possess surface with predominantly negative charges

(Andel man, 1973; Babich and Stotzky, 1983a) • These adsorptil::>n

processes are determined by pH, pE, i9nic strength, concentration

of competing cations, and the concentration and nature of ligand~

present in the medium (Lackie and James, 1974; Farrah and

Pickering, 1977; Gupta and Harrison, 1981 ; Egozy, 1'381 ; Wiley

and Nelson, 1'384). The size of the particulates to which met al

i,ons are adsorbed (grain size) also play an important role

(Snodgrass, 1'380; Forstner, 1982). Mouvet and Bourg (1983)

observed that the complexation of trace metals with dissolved·

ligands does not necessarily prevent their adsorption.

Tada and Suzuki (1982) observed that the adsorption of heavy

metals like Cu, Zn, Cd and Pb can be described by Freundlich

adsorption equation. Gardiner (1974) , and Dudderidge and

18

Wainright (1 '381) reported that the adsorption of trace metals

generally follow Langmuir or Freundlich isotherm. Dudderidge and

0'381) observed in case of four English river Wainright

sediments, in majority of cases Cd, Zn, . Pb, followed Langmuir

isotherm, while in few other. cases, Zn and Pb followed Freundlich

isotherm.

Organic materials both dissGlved and particulate plBY an

important role in the physico-chemical speciation of heavy metal

It is found in geneial that a significant

fraction of dissolved metals are found complexed 'with organic

matter (e.g., Andelman, 1'373; Lerman and Child, 1973; Gardiner,

1'374; Ramamoorthy and Kushner, 1'375; Benes ~t §!!.., 1'376;

Florence, 1977, 1'382; Montgmery and Santiago, 1'378; Van der Berg,

1983; Raspor!ti ~!.. 1984). In aqueous environmental .:;:hemistry

of heavy metals .~rganic materials playa majc.r role by mainly

forming complexes with metal ions, and keeping them in solution.

They are found to mediate large interactions among metal ions.

In case of fulvic acid, one of the important organic complexes

present in natural waters, the main functional groups ()f

5igni ficance to metal ion-fulvic acid interactions are, (a)

phenolic OH- groups which remain un-neutralized except by

prolonged reaction w~th concentrated strong bases; (b) type I

carboxyl groups - those ortho to the phenolic groups, and (c) all

other ionizable function groups like carboxy groups, meta to the

phe-nolic groups (Gamble and Schnitzer, 1973) n The affinity of

19

heavy metals to organic matters are found to follow Irving­

Williams order:

Mg < Ca < Cd < Mn < Co < Zn < Ni < eu < Hg

The heavy metal binding capacity of humic materials ~ere found to

vary in the following order: Soil FA < Soil HA < Peat HA <

Seawater HM < Lake HM River HM < Marine sedimentary FA

< Marine sedimentary HA (Montoura ~~ ~!. 1978).

Binding capacity of organic materials and metals vary ~ith

many factors, the origin of the material being one (e.g.,

Gardiner,

functional

1975) • That may be due to the variation of the

groups in the materials with source and conditions of

formation (Raspor ~~ ~l. 1984).

Ramamoorthy and Kushner (1975) noted different binding

capacity of binding sites in river water with the molecular

weight of the contents.

binding sites to metal

They observed that the affinity of

ion did not alter when treated with

by nitrogen purging and their acidi ficatil:Jn

tleutral i zat i on. It has been noted by them that when total

inorganic carbon was removed the metal binding capacity is not

altered while when total carbon is removed it decreases

significantly to negligible or no-binding capacity.

Organic c,::.mpounds released by human activities like NTA in

natural waters are found to alter the distribution of metal -ions

20

in water (Lerman and Childs, 1973) • It has been found that they

form strong complexes with metal ions like Cu,Pb, Fe, Ni, Co and

Zn, and weaker complexes with metals like Ca and Mg.

Thermodynamic equilibria studies conducted by Lerm~n and Childs

(1973)' show that this type of man-made contaminat1ts bind

virtually with all available Cu, Pb, Ni and Co. After this

metals are exhausted in the medium they bind with Ca and Mg. On

bio~egradation of the~e metal-organic complexes, the natural

equilibria .is restored. Study of biodegradation of these

compounds by Swisher ~~ ~l. (1973) show that these complexes are

fairly easily biodegradable. Stumm and Morgan (1970) state that

the monodentate legand-metal complexes are less stable than

~ultidentate complexes and in case of monodentate complexes the

degree of complexation decreases more strongly with dilution.

With this brief discussion about the speciation of trace

metals in general, the aquous chemistry of each of the metals

studied, viz., Cu, Cd, Ni and Cr, is discussed below.

Cd, an oxophilic and sulfophilic element (Moore and

,Ramamoor t hy t 1 '384) , is a borderline element according to the

classification by Nieboer and Richardson (1980). It undergoes

multiple hydrolysis in pH range observed in the environment. Cd

remains mostly as a divalent species upto a pH of 8. A direct

. relat iOt1 has been obtained by Gardiner ( 1974) between

21

(log (CdOH)"')/(Cd)2+ and pH. Webey and Possel t, (1'374) have

shlDwn that Cd2 +, Cd(OH)(aq) and HCdOz contribute in

increasing ordey to the solubility of the metal in-a carbonate

free Illed i um • At pH ranges below 9 while Cd2 + is the pyedominant

after pH 9 the aqueous hydroxy compounds become

Weber and Posselt (1974) observed that in most

natural waters, Cd solubility is governed by carbonate or

hydroxide

Car-bonate

in systems having low carbonate concentrat ions.

decreases the Cd solubility significantly. For

example, at 0.0 and 5 x 10-4 M total carbonate species at pH 8.3,

the solubility of Cd was 637 mg/l to 0.11 mg/l respectively.

Above pH 9, it is ieported that single solubility calculations

can not be applied because of the formation of hydroxide

complexes. It has been generally reported that higher "hydroxy

species are not relevant in natural systems but in artificial

medium as used in the study reported in this dissertation, where

pH of the medium increases to ~ 9, the hydyoxy compounds become

relevant. Babich and Stotzk~ (1983) have describ~d about similar

si tuat i ot1 where hydroxy groups of the metal can become

significant especially from a toxicity point of view. In system

where chloride content is high, chI oro-compounds of the metal is

dominant (Montoura g1 §l. 1978 - 84% of the total Cd). Ziyino

and YanH)moto ( 1972) also repay ted similai observations in

sea~ater of pH 8.5 (dominant species, CdCI +) •

(1981) making use of single ion activity

22

solubility constants valid at infinity dilution estimated that

.97% of the Cd in seawater is in the form of chloro-compounds.

However, Van der Berg (1983) reports only 12% of total Cd in the

form of CdCI- and 14% in the form of CdCI 2 • Van der Berg

criticiz~s Zirino and Yamamoto's model on its over-complicacy and

states that stoichiometric stability constants valid at the ionic

strength ~f seawater are suitable and simple to use. It is shown

by Hahne and Kroontje (1973) that when the chlor()-CotK~ntration

increases the dominance of chIaro-compounds increase in the order

of CdCI-,

ion. The degree of co-valency of the metal chloride bond was

found to follow the order of Hg > Cd) Pb > Zn, the process of

which can act in mobilizing differ~nt heavy metals from their

sparingly soluble precipitates (Moore and Ramamoorthy,1984).

To organic contends Cd shows moderate affinity. It binds

strongly with sulfhydryl groups. When compard to 6ther heavy

metals like Cu affinity of Cd to organic materials are lesser.

Piotrowicz ~t ~l· (1984) reported that Cd did not

measureably bind like·Cu with marine fulvic acid (MFA) while with

mar ine humi ca.: i d it shows i nt ermedi at e inter act ions. They

explain this situation based on that the Cd interaction may be

related to confoYrfiational considerations such as size rather than

binding through unpair~d electrones. Ramamoorthyand Kushner

(1979) observed the low binding capacity of river water' to Cd and

23

also that it is bound to the fowest molecular weight fractions

(MW < 1400) while Cu shows significant binding with most of the

fractions. Shephard ~~ ~!. (1980) demon~trated the lower binding

capacity of Cd with organic m~tters. In their study of five

lakes of USA it was concluded that in most of the cases Cd 2 + was

the dominant species.

Cu is a borderline element. It shows higher class B

character than Cd (the order is Mn 2 + < Zn 2 + < Fe2 + ~

Cd 2 + < Cu 2'" < Pb 2 +). But Cu+ 1 shows typical class B character

(Nieboer and Richardson, 1'381) • The affinity of Cu for compiex

formation is quite higher than Cd {ref: Irving Williams order).

The metal form complexes with hard bases (Shaw and Brown, 1974)

With ligends like

ethylene diamine it forms typically 4-coordinate complexes. Cu

is seen to bind with particulates and colloids li'ke clays and

hydrous metal hydroxides (Davis and Lackie, 1978; Vu.::eta and

Morgan, 1978). In natural systems with high organic content the

complexation of Cu with these materials are quite high (e.g.,

Gachter ~~ ~!. 1978; Negishi and Matsunaga, 1983) • Lerman and

Childs (1973) demonstrated the higher affinity for binding of Cu ...

to NTA ~nd citric acid at low concentration of the legand. With

concentation increase of the ligand the eu-organic complexes

become predominant. Van der' Berg (1983) observed '331. of the total

Cu in seawater bound to organic ligands while Cd forms l:;tnly 161.

in a system where EDTA were present as Gomple:,;ant. Pamamoorthy

at1d Kushner (1975~ observed that the Cu binding capacity is

mainly confined to 0.45 roM filterable fraction of river water.

The reduction in binding capacity with low molecular weight

fraction ()1400) with removal of organic carbon is also observed.

Blutstein and Smith (1978) report about increase in dissolved

metal cone e'ntr at ion wi th UV i rr adi at i on. They observed an

enhancement in concentrations in an order of two than Cd upon

treatment. Mantoura ~1 El. (1978) emphasizes the affinity of Cu

to humic materials. They observed that in fresh water )90% of Cu

to be bound by humic materials while (11% of other materials are

in bound form. In seawaters as most of humic materials are

co~lexed by metals like Ca, only 10% Cu is in the bound form.

Montgomery and Santiago (1978) documented the affinity of Cu to

organic fractions.

An alteration in the order and percentage precipitation of

eu by fulvic acid with pH is observed by Schnitzer and Kendraff

(1981) • They observed that at pH 7, )95% Cu is bound with

fulvic acid.

~ifl~l

Ni, another borderline element shows class B characters

lesser than Cd.

(Akhmetov, 1'383) ,

+2 state of Ni is the most common state

even though it- can achieve -1 to +4 o:.ddation

states. Ni binds with inorganic legands and forms complexes with

halides, sulfates, phsophates, carbonates and carbonyls. In case

of Ni unlike other heavy metals carbonates precipitation is of-

lesser importance due to the high solubility of Ni carbonates In

natural waters it has been reported that the hydroxide contfol

the- solubility (Reichter and Theis, 1'380). The solubility of

Ni (OH)3(S) are dependant on aging. A decrease in solubility with

aging is reported by the same workers (Active hydroxide

Ks = solubili~y constant). Under

anaerobic conditions the solubility of Ni is controlled by the-

sulfide whidl is comparativ~ly less soluble 10-22. <3).

Morel (1973) using a generalized numerical routine for

equi Ii br i l.lI(1 model showed that 90% of total Ni under o~ddizing

conditions (pE = 12, pH = 7) are in the form of Ni 2 +. For anoxic

conditions (pE = -4; pH = 7) they conclude that )9'3% of-the metal

will be in the form of NiS. Snodgrass (1980) states that in soft

S04 2 -, Cl- and OH- form complexes with Ni while in

fr~sh 'waters (:0 3 - 2 is the pr i nc i pal compl e:-;ant. In seawaters,

40% of the total Ni is in ionic form and 60% in dissolved form,

whi lei n freshwater the percentage b~eak-up is 25% and 35%

the remaining being in adsorb~d form to respect i vel y,

particulates. It is stated by Snodgrass (1980) that unlike

metals like Cu, Zn, Hg, Co, Ag, etc. in case of Ni adsor~tion and

ionic strength are more irflportant than metal and

concentrations in determining the elemental partitioning.

legand

Jenne

(1968) and Singh and Subramanian (1'384) emphasizes the importance

26

of hydrous oxides of Fe and Mn on the control of dissolved

concentrations of many of the heavy metals including Ni.

Reichter and cd-workers demonstrated Ni removal by adsorption

process by various metal oxides i~ aqueous system (Reichter and

Theis, 1980).

Ni form strong complexes with organic legands~ Ni forms

con'lplexes with humic and f,ulvic acids. But the binding capacity

of. Ni to humic acids is comparatively lesser than that of Cu

(Rashid, 1974) and a strong decrease in Ni binding with salinity

is observed CMantoura et al., 1978). Complexation with humic

material.s like fulvic acid in higher ratio (Fulvic acid/Ni > 2)

results in soluble complexes while at lower ratio it results in

insoluble comple~;es (pH 8-'3) (Moore and Ramamoorthy, 1984).

Cr, another metal considered in the study, reported in this

dissertation, hav~ oxidation states from 0 to +6. But among them

the, most common

(Akhmetov, 1983).

form is Cr-3 form followed by Cr-e form

Coordination number of six (for Cr-3 ) and four

(for Cr-S) are mostly typical of the element. Unlike all the

three metals considered above,

is more an hard acid (Pearson,

Cr in the most common form (Cr+3 )

1963) . Cr+& compounds are strong

oxidants and are converted to derivatives of Cr+3 in redox

reactions. In natural medium they yield Cr+3 hydroxides

(Cr(OH)3), in acid medium cation tomplexes (Cr(OHz)e)+& and in

27

alkaline me~ium anion complexes (CrCOH)e)3- (Akhmetov, 1'383).

Cr+3 is found to form hydroxo and oxobridged polynuclear

complexes (Stumm and Morgan~ 1970; Moore and Ramamoorthy, 1984).

They form strong compl exes wi th oy;ygen, ni trogen and su1.fur donor

atoms. In reducing conditions Cr+e is converted to Cr+3 and in

well oxygenated water Cr+6 is thermodynamically stable species.

Hexavalent species is easily reduced to the trivalent form by

ferrc.lus ions, dissolved sulfides and organic legands with

sulphydryl groups. Benes ~i ~!. (1976) reported that the

trivalent form of this metal is generally bound to the colloids

and complexes of M. wt. higher than 10,000. They observed that

around 881. of Cr can be retained by dialysis membrane. Pfeiffer

( 1980, 1982) discuss about the conversion of hexa~form to

tri-form· in a river system. During the transport the hexavalent

Cr released to the river partially got converted to trivalent

form. Pfeiffer and his co-workers discussed the affinity of the

lower oxidized form to partic~lat~ matters. As the hexavalent

form of the metal is typically anionic it shows little affinity

to organic ligands (Frey ~i ~!. 1983).

IQ~!~lIY

Among heavy metals many are inevitable to organisms as their

role as micro-nutrients (e.g., Bowen, 1 '366) • Heavy metals like

Cr, Co, Cu, Ni, Mn, Zn, V, etc. are recognized to have important

Tole in the metabolic activities. Cu is known to function as

28

micronutrient to algae as well as higher plant~ (Meyers, 1962;

Manahan and Smith, 1973; NAS, 1975; HWl t sman an d Sund a, 1 '380) •

It is a constituent of plastocyanin, a protein involved in

photosynthe~ic electron transport (Katoh ~~ ~l. 1'352; Lehninger r

1'382) • Bowen (1972) reported }30 proteins and enzymes which

contain Cu. Hemocuprein or superoxide dismutage, cytol:hrome C,

ascorbic' acid oxi dase, cytochrome oxidase, . tyrosinase,

hemocyaninin etc. are some of the Cu containing enzymes/proteins

(Bol i n r 1 981) • Ni also pI aysi mpor tatlt role in fflet abol i I:

activities of microbes and plants (Bertrand and DeWolfe, 1'367;

1975; Polacca, 1977; Van Baalon and O'Donnel, 1978;

Kaltwasser and Frings, 1980; Daday and Smith, 1983; Oliviera

and Antia, 1984 ). The effects of Ni difficiency is reviewed by

Kirchgessner and Schnegg (1980). In case of Cr the trivalent form

is known to have important role as a trace nutrient, while Cr+6

is highly toxic. Mertz (1'359) reviewed the physiological role of ,

Cr with the emphasize of mammalian system. He discussed about

• the possible role of the metal to bacterial system also. The

possible defficiency problems arising from the lack of Cr is also

elaborated 1n the same review. Cr is required in many

enzymes/factors like phosphoglucomutase, trypsine,

tolerance factor etc. Cr defficiency alter the fermentation

capacity of many bacteria. Unlike the metals discussed above, Cd

is a non-essential r non-beneficial metal to the biological syst~m

(NAS, 1972; Jenkins, 1980; Hardy ~t ~l. 1984). ,

29

The toxicity of heavy metals are influenced by many external

and i nt ernal factors. Among the external factors, pH of the

mediu~, temperature hardness, dissolved oxygen, presence of other

metallic cations, organic compounds etc. are included while the

internal factors comprise of sex, stage in life cycle, body size,

trophic status, species specific tolerance etc. ,These factors

influence the toxicity by their effects on the bioavailability of

metal ions as well as by their influence on the physiological

conditions of the organisms. Dc dl i a i (1977 ) has identified

t.:)xicity mechanism of metal ions into three categories: (1)

blocking of essential functional groups of biomolecules; (2)

displacing the essential metal in biomolecules, and (3) modifying ..

the active conformation of biomolecules. The toxicity of heavy

metals are closely related to their physiochemical features (like

formal charge and ionic radius) which determine the affinity of

specific binding sites to biological macromolecules. Most ()f

these elements show high affinity to biologically significant

functional groups like sulphydryl, amino, imino, carboxyl etc.

and bind with ~hem leading to 'denaturation' of the concerned

protein/enzyme with respect to its specific biological functions.

Vallee and Ulmer (1972) demonstrated the affinity of heavy metals

like Cd to sulfhydril groups. Fisher and Jones (1 '381)

correlation between the toxicities and solubilities of metal

suI fides. They proposed the binding of heavy metals to the

'sulfhydril' groups as one of the mechanism of metal toxicity.

30

The possibility of similar process of metal binding to similar

groups of especially non-critical proteins has been discussed by

Rothstein (195'3) , F.~achl itl ~i e.l.. (1'382) , as a process of

deto:d fi .:at i on. The ~ffinity of metals to biriding sites are

generally found to follow Irving Williams order (Irving and

Williams, 1953) • Shaw 0'3(1) later demonstrated that the

toxicity of transition ~etal ions on living organisms as well as

enzymes followed similar order. In the process of Isomorphous

interchange, another possible mechanism leading to metal toxicity

(Hanzlik, 1981 ; Occhi ai, 1977). The structure of the competing

atoms are important because the replacement of one metal

(essential element) by another (toxicant) is dependent on the

structure recognition by the binding sites on the protein moiety.

Substantial reduction or alteration in the activity of metallo-

enzymes by similar substitution is gene~ally observed (e.g.,

Springgate ~i ~.l.. 1973; Eichorn, 1975) • The structural

similarity between the toxic metal and an importatlt physiological

substrate is one process which leads to the entry of the for~er

into the cell (e.g., methyl mercury derivatives of cystine,

homocysteine and amino acid transport, chromate ion and active

transport system for sulfate ion, Ti+ ion and potassium

Hanz 1 i k, 1'381 ) • Many of the toxic metals are highly dependent on

the nutrient elements like Fe in their transport to the cell

(e.g., Harrison" and Morel, 1983). In these cases, they interfere

with the nutrient uptake (Goering ~i e.l.. 1977; Huntsman and

31

Sunda, 1980; Rueter- and Morel, 1981; Rueter ~t ~!. 1981; Foster

and Mo~el, Sunda and Huntsman, 1983; Harrison and Morel,

1983), and result in difficiency of nutrients. Alteration of the

permeability is affected (Hassel, 1963; Steeman-Nielsen ~t ~!.

1969; Sheih and Barber, 1973l,by some metals like Hg and it

results in interference of vital activities like Na, K transport

of cellular system.

Cd which is not so far known to have any biological role is

highly toxic to organisms. Its toxicity to different groups of

org~nisms have been reviewed by Flick ~t ~!. (1971), Friberg ~~

al. (1971), Fishbein (1974), Venugopal and Luckey (1978),

Nomiyama (1980), Forstner (1980). This metal shows high affinity

I to sulphydril groups (Vallee and Ulmer, 1972) and also competes

with many metals of biochemical importance like Zn, Fe, Ca and

Mg. It is also found to bind with nu~leic acid through the

phosphate group at the ribose phosphate backbone (Eichorn ~i ~i.

1970). A number of studies have been reported on the toxicity of

Cd on different algae in laboratory c4ltures as well as in the

field. Main parameters considered in the studies were growth and

survival (e.g., Vocke, 1978; Fisher and Frood, 1980) effect on

morphology (e.g., Adshead-Simonsen ~t ~!. 1981; Rachlin ~t ~!.

1984), effect on biochemical activities (e.g., Pietilinen, 1975;

Delmotte, 1980; DeFilppis ~t ~!. 1981; Irmer ~t ~i., 1983) etc.

for single species as well as mixed population studies.

32

The study conducted by Vocke (1978) demonstrated the

di fferential

AAM medium.

t6xicity shown by four fresh water algae grown in

He observed that 0.01 ppm Cd was necessary to

while comparable inhibition was observed in

case of §£1~!J92:t.r.!:!ffi SPa with only 0.05 ppm Cd. At 0.3 ppm Cd all

the experimental species showed algicidal or algistatic response.

In the study it was found that Cd is the most toxic metal to all

the organi SIT,s. The toxicity order in terms of EC50 was for

§f~~~~§ffiY§ Cd > As > Hg > Be;

§Ql~!J~§tr.!:!m Cd > Hg > Be > As and ~ifr.QfQ!~y§Cd > Hg > Sen Of

the four species studied by Vocke (1978), §~~n~~~§mY§ Spa was the

least tolerant. Bartlett Q.:t. Sll· (1974) studied the inhibitory

effect of Cd on §~l~nSl§tr.Ym £~Rr.i£Qr.nYtYm using the algal assay

procedure bottle test (EPA, 1'371) • They observed that 0.65 ppm

of the metal as algicidal to the species. Cd at a concentration

of 0.05 ppm initiated inhibition of the growth of the alga and

0.8. ppm completely inhibited the process.

extension of the log phase by Cd exposure.

They observed the

Hart and Scaife

(1977) showed the gradual inhibition ~f the growth of ~hlQr.~ll~

I!~re!:lQi~Q§2 with 0.25; 0.50 and 1.0 ppm at"pH 7-8. An increase

in doubling time from 11 hrs to 21 hrs was induced by the

addition of 0.25 ppm Cd. A slight reduction in toxicity was

noted with an increase in the pH from 7-8. The doubling time was

increased to 16 hrs at pH 8 with 0.2 ppm mg/l of the metal. The

33

alterations in toxicity of Cd due to pH changes to different

organisms is widely observed and the findings are as wide as the .

metal used and the experimental species. Babich and Stotzky

(1977a, b; 1'380, 1'383) demonstrated the situation in different

species of bacteria and fungi. The enhanced toxicity observed by

the workers (1'377, 1'380, 1983) is explained that the hydroxy form

of Cd, viz., Cd OH+ is capable of penetrating more the biological

meo".brane for the decreased competi t i on between H+ to Cd 2 + /Cd OH-

ions at the binding sites due to the high pH. But genBrally~

Cd~'" is the dominant species at pH > 8 (Weber and Posselt, 1974)

thus indicating that in case of Hart and Scaife (1977) study the

toxicity is mainly due to the ionic form.

Rachlin ~t !§!l· (1982a) observed a toxicity order of Cd > Cu

A decrease in the growth

rate (f() and a corresponding increase in thB number of

days/division [T(d)] is observed with increase in metal

c.::.ncentr ation. In another study Rachlin ~t ~!. (1982b) noted a

based on the EC50 values 96 hrs. it is noted by the authors that

though (b.!.Qr..~l.!.~ ~~£.£.b.£r..QQ.hi.!.~ have lowest EC50 96 hrs its

response was flatter with increasing concentrations of the metals

than the other two experimental species indicating a wider range

of tolerance shown by the green algae over the critical

concentration range. With the similarities observed in the

34

response of taxonomically more tlosely related species; Rachlin

(1982b) explored the possibility of an organizational 1

framework for physiological responses which can be used in

predicting the physiological response of different organisms from

the knowledge of their physiological relations. In another

study, the same authors (Rachlin §1 ~l. 1983) report that

Ne.':~!'i£!dl~ in£§r.1e re-sponds to different ITletals in the following

toxicity order: Cd > Pb > Zn > Cu. Rachlin ~1 ~l· ( 1983)

conducted the study in LDM medium free of chelators (starr,

1'378) • The EC50, 96 hrs reported from this study was 3.01 +

0.011 ppm Cd (26.8 uM).

~b!Qr.§l!~ species were studied with respect to Cd toxicity

by workers like Jong (1965), Hart (1975), Hutchinson and Stockes

(1975), and Rosko and Rachlin (1977). Jong (1965) demonstrated

that £;blQ!.§lle y!:!1ge!.i§ could to::llerate a concentrati'::ln of uptl::l

0.09 ppm Cd without any effect on the growth. It was seen that

the. lowest concentration which prevents growth of the species was

.::: 0.14 ppm. .::: 0.25 ppm of the metal was found to be inhibiting

to ~b!Q!.§lle Q~!.§DQi~Q§e during the logarithmic growth phase by

Hart (1975) • Hutchinson and Stockes (1975) noted inhibition of

growth of Chi orella by Cd at a concentration of 0.05 ppm. Fisher

and Jones 0'381), for 8§1~r.iQD~11~ jeQQni£~ determined a tmdcity

hierarchy of Cu > Zn » Pb > Cd. Conway (1978) and Conway and

Williams (1978) reported a linear decrease of growth rate with

0.002 to 0.00'3 ppm Cd concentration and complete inhibition by ::

Conway and Williams (1978)

noted that 0.01 ppm Cd which totally inhibit growth of

~§ :t~L!.'~!:l ~!.!.£ Spa dOl?s not affl?ct thl? gY"owth of [r:.~gi!.~r:.i~.

£:r:.Q~~!:l§i§· Ll?s and Walkl?r (1984) observed the i nt I?rml?di at I?

t'jxici ty of Cd to Cu and Zn to the expl?Y"iml?ntal spl?cies

A concl?ntration of 1.0 ppm Cd was found to bl?

necessaY"y to inhibit thl? growth rate of algal? significantly While

comparable rl?duction was inducl?d by 0.2 ppm of Cu or 2.0 ppm of

Zn.

For cyanophyte and

Rachlin i:i 21· (1984) dl?tl?rminl?d an EC50 96 hrs

(incipient ll?thal concentration) of 0.118 ± 0.04 uM (0.013 +

0.003 ppm) and 0.11 uM (0.012 ppm) rl?spl?ctively. Thl?Y found that

this specil?s were morl? sl?nsitivl? than eukaryotic algae eaY"lil?Y"

studil?d by thl?m, viz.

';'!.Q.§t~r.i\:!ffi and ~~y.!.£,!:!!.~. !.!:l£.~r.t£. - An inc Y" I?ase c' f numbl?Y" 0 f days/

divisi':>n (T<~~) from 1.95 to 4.57 was also seen with incY"l?asl? of

metal from 0.0 uM to 0.4359 uM (0.05 ppm). In case of a spl?cil?s

of blul? grl?l?n alga, Laube i:t ~!.. ( 1980) found that total

inhibition was affected by 1-11 ppm Cd while Stratton and Corke

(1979) observed similar effect with 0.06 ppm ml?tal.

A significant inhibitory effect of 0.006 ppm Cd and sl?vere

inhibition with 0.061 ppm Cd had bel?n notl?d by Klass gi £1.

Rosko and Rachlin (1977) ,

,

Fisher and Frood (1980), Lue-Ki m ~t ~!... ( 1'380) , Hac hi in !"d.. ~!..

(1982, 1983) observed that cell division is one of the most

sensitive parameter to Cd toxicity. These observations are

significant in view of DeFillippis ~t ~!... (1981) findings that in

case of g~g!.~n~ Spa heavy metals like Cd, Zn and Hg inhibit the

NADP-oxidoreductase, possibly by binding the sulfhydril groups by

the nletal. The inactivation of the enzymes results in the short

supply of NADPH to the cells.

Dongmann and Nurnberg (1982) studied the effect of Cd on a

chain forming marine diatom by considering generation time~ cell

density and chain length as itidicators. The

e:,;per imental

It was found for the species that upto 25 u mole (2.81 ppm) the

mean generation time remained more or less same. And only the

generation time rose from 24 hrs to 28 hrs, when

concentration of Cd reached about 50 u mole (5.62 ppm) •

the

The

experiment was conducted in enriched seawater medium. A decrease

in chain length was also observed with increase in Cd

concentration. Among the three toxicity indicator parameters

studi ed, Dongmant1 and Nurnberg (1982) conc I uded that cell densi t y

is fIlore- se-nsitive than mean chain length or growth rate. The

estimation of generation time as the indicator the value was: 90

uM (3.4 ppm) and in case of chain length it came to 15 uM (1.7

ppm) and for cell density. the response was in the level of 1-10

uM (0.1 - 1.1 ppm).

37

Oeviprasaq and Oeviprasad (1'382) .:onduct€.'d toxicity studies

of Cd, Pb and Ni on three species of fresh~ater algae, viz.

Spa It is found in their studies of, the three metals that Cd is

more toxic than the other two.

!~lc~:tl::!~ 5.0 ppm caused complete death of the organism while fQr

10.0 ppm was necessary fQr similar effect.

Their medium of experiment was Chu-l0 (Stein, 1'373).

Canton and Slooff (1982) conducted Cd tQxi~ity studies with

organisms of different trophic levels and observed toxicity in

the decreasing order of Q§Qb~Di~L Q~~~i~§L XgDQQ~~L ~b!Q~~ll~L

Br§£b~~~niQL §~!mQDgll§ and EQg£i!i~. They determined a nontoxic

effective level fQr ~b!Qrgl!§ ~y!g~ri§ as 2.6 ppm Cd 2 + for 48 hrs

and 72 hrs and 1.5 ppm for 96 hrs. They Qbserved EC50 for the

species for the same time intervals were 5.1,' 4.4 and 3.7 ppm

respectively taking growth inhibitiQn as the toxicity parameter.

Bentley-Mowat and Reid (1977) engaged representative species

of four groups of marine phytQplanktQn tQ study the effect Qf Cd,

Cu and Pb. The species selected were Ig:t~~~~!rni~ Spa

(Prasinophycaea),

Eb.g:QQ.§.t;.:t:i.!!drn

Q!dtli'!!ig:!!.i'!

tr..i..t;.Q!:.tll:!t!dIT!

(Chlor':Jphyceae-) ,

and

(Bacillariophyceae) and ~r..i..£.Q2.Qb.§.~r..~ g:;LQ!:l9.§.ti'! (Heptophyceae-).

The algae were cultured in 5-88 medium (Droop, 1968) at 167.

salinity. ~~i£..Q2Qbgr..~ Spa was fQund·to be the most sensitive t'a

38

Cd and shown lowest growth • It was seen in case of of

. I~.t[~§~!.mi2.t.. Eb~~f!Q~f.t.:t1!Jm an d h[ifQ§Qb~~r.~, Cd is more toxic

than Cu or Pb and for Q~n~!i~!l~ there was no difference in

toxicity between the metals.

studied by H,~l1 ibaugh ft.t. 91. ( 1'380) • The experiment, showed for

natural population the following order Hg » Cd > Pb > As > V

> Zn > Cd = Ni > Cr Sb Se As I I I. They observed a slight

depression in growth of Ib~1~2§!Q§ir.~ SPa cultured in enriched

seawater with 500 nM (0.056 ppm) Cd while more or less equal

respl:>nse was obtained by 10 nM (2 x 10-:5 ppm) of Hg. Ber I atld ~t

(1977) determined that about 440 nM (0.05 ppm) of Cd was

necessary to show inhibition of growth in case of §t~!.~iQ'1ft!:f.!f!

In a study conducted on natural phytoplankton

population, CQok (1975) observed that Cd upto 0.124 ppm did not

alter growth. The upper range of tolerance shown by the same

popUlation was ~ 11.24 ppm.

Hart and Scaife (1977) observed alteration i tl the

mol'" phol ogi cal appearance of hb!Qr.~!.!~ Spa due to Cd exposure.

The organisms formed aggregate of 4-6 cells. The aggregation was

tll:>t detachable with homogenization or detergent treatment. They

assume that this morphological change, resulted possibly due to

the failure of the parental cell wall to disintegrate after

sporul at i on. Rachli n (1984) utilized morphometric

analysis as one of the measure of toxicity. A si gni ficant

39

reduction in cell size in case of en~~~~n~ i!Q§=~g~~~ was induced

by Cd e!l;posure. The lowest concentration to have any effect on

cell dimension was 11.83 uM (1.3 ppm) . Reduction in the surface .

area of cell thylakoids~ and intra-thytakoid spaces was observed

in case of the metal exposure. Reduction in the volume of

polyphosphate bodies, increase in the nuniber and volume of lipid

i ncl usi cons, and the number of cyanophycin granules and shrinking

away of plasmamembrane from the cell wall were other structural

alterations induced by Cd exposure. An alteration in this

configuration of the cells of a fresh water diatom I~~~!!~~i~

f!9.S.S.!::!.i.2§.€! has been report ed by Adshead-Si monsen f:! ~!.. (1'381) on

Cd exposure. On addition of 0.001 ppm metal the diatom changed

to a straight configuration rather than its normal zig zag

arrangement of cells .. With the application of Cd . in

concentrations of 0.03 to 0.1 ppm, ultra structural changes were

1976) • An increased

number of zoosporangia with Cd exposure was observed in case of

(1'383) • They also

noted the disarrangement of thylakoid systems of chloroplasts and

development of fingerprint like structure on treatment with 5 uM

(0.56 ppm) metal.

Cd inhibit many biochemical activities like photosynthesis,

nitrogen fi xat i on, nutrient uptake, etc. of the algae. On

40

in photosynthesi s and chlorophyll a content was noted by IrrJler ~t

(1983) • The effect was found to be enhanced by both

concentration of metal added and duration of exposure. With the

addition of 1 uM (0.1 ppm) Cd the chlorophyll content got reduced

by 33% with 48 hr exposure while 3 hr exposure resulted in 3%

re-duction. In presence of 20 mM (2.2 ppm) metal, the reduction

was 89% and 9% with 48 hr and 3 hr exposure respectively. 1 uM

(0.1 ppm) Cd reduced the photosynthetic oxygen production by 13%

by 3 hrs and 23% by 24 hrsa In case of 20 uM (2.2 ppm) the

reduction were 84% and 100% respectively.

Decrease in protein and chlorophyll content in presence of

Cd was reported by Hart and Scaife (1977) in case of ~b!QL~!!~

Lehman and Vas Concelos (1979) observed inhibition

of photosytlthesi sand respi r at ion 0 f marine di atom ~!in9.LQ:tb~£.£.~

£1.Q§igr.i!:!!!! at 0.001 ppm Cd. )50 uM (5.6 ppm) metal was found to

inhibit the chlorophyll synthesis in hblQL~ll.e QYL§-!:JQ.!.QQ§.e (Lue-

Ki,.n ~:t ~l· 1980) •

decrease in chlorophyll content with Cd was noted by Rebhum and

Ben-Amotz (1984). One pe~uliarity observed in the experiment was

that when at the level of 0.5 to 3.0 ppm Cd, the fall in

chlorophyll content was sharp while in the range of 3-10 ppm

the effect was relatively mild and after 5 ppm reduction with

further concentration increase became negligible. Overnell

(1975) determined a toxicity hierarchy based on photosynthetic

41

oxygen evolution for Q!:!t!~li~ll~ 1~r.1i91~f.1~ as Cu :: Hg » Pb > Cd

and Eb~f.QQ.e.f.1~1!dill t.r.if.9r.f!!]h\t.!df!:! as Hg = Cu » Pb > Cd. Hongve ~:!!

( 1980) found an intermediate level of inhibition of

photosynthesis of natural phytoplankton population.

order of toxicity was Hg > Cu > Cd > Pb ~ Zn.

The observed

Inhibition of nitrogen fixation by Cd was demonstrated by

Henriksori and Daselva (1978). On exposure of Nostoc Spa to Cd at

concentration of 0.025 - 0.125 ppm the process was inhibited. In

Delmotte (1980) reported inhibition

of photosynthesis by 1.9 ppm Cd and nitrogen fixation by 2.0 ppm.

Conway (1978) demonstrated the influence of Cd on N0 3 -

The alteration in N03 -

metabolism of Ib~lf!§§iQ§ir.~ 11~Yi~:!!ili§ was observed by Li

(1978) • The study shown at low N03 concentration, concentration-

dependent severity of Cd toxicity to the experirilental spec i es.

Recent 1 y, Harrison and Morel (1983) reported the influence of Cd

on Fe uptake and vice versa by Ihf!lf!§§iQ§ir.~ ~~i§§ilQgii. At low

concentration of ferric ion ~ simultaneous decrease in growth and

Fe accumulation was observed. Lewitl (1954) reported the

inhibition of silicic acid uptake of diatoms by Cd.

On a natural popUlation of §Qir~liD~ Ql~:!!~D§i§ collected

from a soda lake and grown in the same water Kallqvist and

~2

Meadows (1978) f6und that Cu addition upto 0.02 ppm reduced the

growth rate significantly (control growth rate 0.19 and-with 0.02

ppm Cu, 0.12) when Cu concentration was increased to 0.2 - 2.0

ppm the number of trichomes decreased to less than initial number

within three days. A sequential reduction in growth rate

(divi~ions/day) was observed by Sunda and Guillard (1976) upon

addition of eu to estuarine diatom Ib~!~§§iQ§i~~ Q§~~~Qn~n~ and

green alga ~~nnQ£bri§i§ ~iQm~§ culture grown in enriched

seawater. In case of diatom, growth was inhibited by a

concentration of 3 x 10-1~ M (0.019 ppm) and complete growth

inhibition by 5 x 10-9 M (0.32 ppm) of Cu. Partial growth-

inhibition of ~~nnQf.bri§i§ EiQm~§ was observed in the activity

range of 4 x 10-11 to 2 X 10-9 M Cu. They observed a direct

correlation between growth inhibition and Cu+2 activity rather

than total Cu added. The i~fluence of pH and Cu complexation is

also demonstrated in the study. Jensen ~t ~!. (1976) studied eu

tolerance to marine diatoms, viZa

Ib.~!s!22i9.2i.r.S! Q.2!:±~9.Qtls!!:ls! and Eb.S!~Q~S!S.tl::!!:±ffi tr.i.s.Q~!:l!:±t!:!ffi grown in

dialysis and batch cultures •. Of the three algae engaged in th~

exp~r i ment, was found to be the most sensitive

followed by Ib~!E§§iQ§i~~ Spa The

concentration which induced growth-reduction was 0.01, 0.025 and

0.4 ppm of the metal respectively_ A trend of enhanced

accumulation with the sensitivity can be observed from their

e:t:per i ment. Ey"ickson ~i E!' (1'370) found an order of sensitivity

43

to Cu among six sp@ci@s of the algae as ~mQhi~ini~m

Of th@s@ si~;

species the first three exhibited 80% reduction in growth with

0.05 ppm Cu, the fourth one showed ~ 36% r@duction with 0.1 ppm;

fifth showed total inhibition with 0.15 ppm and sixth one show@d

50 per cent r@duction with only 0.4 ppm.

(Steeman-Nielsen and Kamp-Nielsen, 1970) 0.001 and 0.005 ppm Cu

resulted in 24 and 48 hrs lag in growth. Lat@r the alga resumed

the growth rate of control. R@duction in growth rate with Cu

addition was observ@d by Bartl@tt@ ~t ~l. (1974) in case of

Bentl@y-Mowat and R@id (1977) found

that E'b~~.QQ~f..t.:il~!!} .t.r.if.2r.!}~.t.~!!} and hr.if.Q§Q.h~~r.~ ~lQ!:Jg~:t:.E! survive

In continuous culture with no diminution of growth rat@ with the

addition of 10-3 M and 10-4 M Cu, respectively_

Under reverin@ conditions, Vlotz (1981) r@ported growth

r@duction due to Cu. The alga@ studied w@r@ h.hl~m:i~2mQtl§!§ sp.,

The experiments were conduct@d in dialysis tubings. Diff@rence

in between isolates from different sites'was obs@yv@d -possibly

because of the development of Y@sistence as observed by Fost@r

(1977) , Hall (197'3) , Sh@hata and Whitton (1 '382) and

Whi tton and Shehata (1'382),

With 0.4 ppm Cu

the growth was compl@t~ly inhibited in

laboratory experiments. A variation in toxicity due to

temperature changes, especially with high conc@ntraticin of the

metal (0.25 ppm) was also noted. In the study conducted by Klotz

(1981) a shift from optimum temperature increased copper to:dcity

Cairns ~t ~l. (1978) have stated that at optimum temperature the

was higher. Comparabl@

observations were reported by Sharma (1985) on ~Q£:£:i2ii2 Spa and

§Q!.r.~!.iQ£: sp. '

Growth of eD~~~~D~ ~~~i~~ili2 was reported as completely

inhibited due to the addition of 30 ppm (Young and Lisk, 1'372) •

They also found that g~@@n alga to be more resistant than blue

gr e@t1s. Steeman-Nielsen and Wium-Anders@n ( 1971) observed

complete inhibition of growth of the diatom ~ii~2£hi~ Q~!.~~ grown

in Oste,rland medium B with addition of 7.5 to 12.5 ppm Cu. An

increase in cell density, increased the required concentration to

inhibit growth. A lower initial cell density (2 x 10-~ cells/I)

required lesser conce,ntrations (7.5 ppm) than a higher initial

cell density (10 7 cells)(12.5 ppm). The cell number was found to

be of importance in eu toxicity by Young and Lisk and Steell'lan-

Nielsen and. Kamp-Nielsen 0'370) also.

density the toxicity become higher.

Generally, with less cell

"5 't'

An enhanced toxicity of Cu due to the reduction in pH has

been reported by Michnowicz and Weaks (1984 )

At pH 10, when the species showed highest growth,

the metal, with a concentration of 2.0 ppm waS showing higher dry

weight biomass than at pH 4 with equal metal concentration.

In a holistic study of Cu 2 + stress on an aquatic microcosm

Suguira ( 1'382) demonstrated the role of species

interactions and system stabilization on the toxicity. The

variation in toxicity with successive stages has been elaborated

by Suguira ~E ~1. (1982). On 0.7 ppm metal addition at the

beginning of the culture,

g!.igQhb.~S!i~§., and the rotifers were E-liminated but when the metal

was added after a few days similar catastrophic changes were not

observed. A reduction in population was affected, however, the

algal groups showed viable growth rate.

Les and Walker (1984) studied.the toxicity of Cu to fresh-

water blue green alga ~bLQ££Q££Y§ Spa Lowest concentration

which showed detectible toxic effect wa-s 0.1 ppm. Rachlin ~!:. 91.

( 1983) for Ney:ihY!.9 ilJ.h~r..!:.9. determined EC50 '36 hr as 164.5 um

(10.45 ppm) based on growth studies and 5 uM was estimated as

EC50 96 hrs value for ~b.!.Qr.~!.!.9 §'9£'hb:~r.QQ.b.!J_9 (Rachlin ~i 91.

1982) _ Peterson (1982) determined an EC50 (50 per cent reduction

in growth rate) for §h~IJ.~~~§.mY§. gY9~r.i£.9Y~~ as 10-s . B M Cu· (aq).

A direct relation of the growth rate and ionic metal

concentration was observed in this exp~riment. A relationship to

predict the growth rate from the concentration was proposed by

Peterson (1982) in this study.

Morphological changes due to Cu exposure were reported in

case .::>f many al gal spec i es. Massalski ~i ~l. (1981) noted ultra

The- gr eetl

alga produced multinuclear gaint cells with thickened walls when

exposed to 10-4 M Cu (6.35 ppm). Pac hi i t1 ~:t ~!.. ( 1'382c) used

rfIc'rphometric analysis in evaluating reSpOtlSes of E!.~£.iQ'2~!I!2:

t!Qr..Y~t:!!::!!I! (Cyanophyceae) on e:-;posure to eight heavy metals, Nn,

Ztl, Hg, Cd, Ni, Co, Zn, Ag and Pb. In case of El~£.:tQn~!I!~ unlike

that of !:!lJ.~Q.~~'2~ (Padllin ~:t ~!.. 1'384) no change in cell size due

to Cd was noted. eu and Pb produced increase in cell sl .. ze.

Reduction in cellular lipid was also induced by Cu. Cu induced

only a slight decrease in thylakoid surface area and Co, higbest

increase. The absence of an~ cellular distortions after exposure

to 100 ppm Cu for 4 hrs show that the metal induces synthesis of

an active cellular material including the cell wall or a

depolymerisation of the mucopolymers in the wall matrix to effect

in stretching of the cell wall thus accofYIlt"lodating a net increase

in the cell size. Cellular distortion with Cu exposure was

observed by Sunda atld 13uillard 0'376) in .:ase of Itl~!.£iaaiQ2.ir:.~

I!§eu!1QM~ •. Cellular elongation and morphological distortions

was associated with pH in range 8.6 - 8.3 in the medium. OSft'lotic

disorganization and swelling of the cells content was induced by

47

10-4 M Cu to Ditylum cells (Bentley-Mowat and Peid, .1'377).

Hollibaugh ~1 ~!. (1980) noted, out of column cells and clumbs of

daughter cells that had failed to separate after division when

exposed to 100 nM Cu in case of Ib~!~§§Q§i~~ ~~§1i~~!!i§.

Reduction in chlorophyll and primary production by §Qi~!dli!:l§!

l!!.at~!:l§i2 by Cu addition was reported by Kallqvist and Meadows

(1978) • 70% reduction in chlorophyll a after 8 days exposure and

80% reduction in primary productivity was resulted by addition of

2.0 ppm of the metal.

0.1 to 0.5 ppm Cu.

Productivity showed sharp reduction from

But further increase in Cu concentration

l~esulted in ,decrease of inhibi tc.ry ef fect. Photosynthesis by

.Qb!.Q~~!!.9 QYr:.~!'}Qif!Q§9 was reduced to 50% by 4. 8 ~; 10-7 M (0.03

ppm) Cu (Steeman-Nielsen ~1 ~! 1'369). For the diatom ~i1~2fbi~

Q~!~~ more or less similar response was obtained with '3.5 x 10-e

M (0.006 ppm) Cu (Steeman-Nielsen and Wium-Andersen, 1971).

Saifulla (1978) reported the reduction in photosynthesis by

marine dino-flagellate t by 0.005 ppm Cu.

Decrease in carbon assimilation (C14) was reported in case of

(Trichodesmium) by Peuter (197'3) •

Difference between surface organisms (collected from surface) and

25 m depth organisms was noted in case of sensitivity_ In case

of depth organisms, 10-10 M Cu inhibited carbon assimilation by

50% while for su"rface organism the same effect was ellicited by

I

48

10-9 • 3 M CU. Steeman-Nielsen and Wium-Andersen (1971) noted that

in case of diatoms the response to the metal was very fast.

Photosynthesis showed inhibition within a short time by ~h!QL~!!~

Q~L~nQiQQ2~'

photosynthesis.

Illumirtation rate affected eu inhibition of

Studies have shown that in case of long term

exposures green algae show higher sensitivity while in case of

short term primary productivity studies diatoms are more

sensitive. Wu and Lorenzen (1984) showed that photosystem II is

more sensitive to Cu, than photosystem I from their studies on

With light and dark change during a light period the

sensitivity of O2 evolution to CU2~ is found to fluctuate.

Davies and Sleep (1980) observed that photosynthesis of a

coastal marine plankton assemblage gets inhibited by 0.001-0.0025

ppm. The influence of Cu on nitrogen fixation was studied by

Horne and Goldman (1974). They reported that 0.005 ppm of the

metal resulted in the reduction cif nitrogeh fixation by blue

green algae in an eutrophicated lake.

eu inte~feres with nutrient uptake mechanism.

uptake is inhibited in an irreversible way by Cu.

Nitrate

Further.

supplimentation of the nutrient does not make the resumption of

the process (Harrison ~i ~l· 1977). The mechanism of the

interference may be due to the inactivation of ATPase enzyme

required for nitrate uptake (Kanazawa and Kanazawa,

Uptake of silica is also inhibited by Cu (Goering ~i ~l.

1969).

1977;

49

Reuter gt. ~1· 1981) • Cu inhibition of silicic acid uptake was

"i mmedi ate. In case of the study conducted by Reuter ~t. ~1.

(1981 ) on (Ibe1222iQsir.~ sp.) the inhibition of the nutrient was

very sharp -within one hour

between Cu and silicic acid,

itself. To explain antagonism

they hypothesize a silicic acid

transport site 'which gets inhibited by Cu 2 - and serves as a

transport site for the metal. Silicic acid uptake inhibition by

eu was reversible (Goering ~t. ~l. 1'377 ; F.:euter ~t ~l· 1'381) .

Mol" el (1'378) concluded that the interaction betw"een

silicates and Cu is of purely physiological nature.

~i£lgl

Growth inhibition by Ni was studied widely on micro algae

(St okes ~.! 21. 1973; Upitis g.! §1. 1974; Skaar ~:t. 91. 1974;

Hutchinson and Stokes, 1975; Stokes, 1975; Patrick g.! ~l. 1975;

Fezy ~t i!1. 1'379; Hollibaugh gt ~l· 1980; Dongmann and Nurnberg,

1982). A number of different algal species were engaged in

growth studies (Spencer, 1'380). SpetKer (1980) I ists around 2(Y

species of algae used by various workers.

(-1979) studied the growth of a fresh-water

diatom N~Yi£~12 Q~!!ifY!Q§~ under Ni stress. Concentrations of

1. 7 :,.; 10-e. M ,(0.1 ppm) reduced the population growth rate by 501.

They assume that the toxicity observed in this case is

exclusively due to the ionic Ni present in the medium. Wi th O. 1

ppm the doubling time got increased by a factor of 1.5 of the

50

cOl1trol. Spencer and Green (1981) reported on the toxicity of Ni

to seven species. The species engaged in the study we~e ~Q~~~~Q~

R~lated species showing close

range toxicity was observed in the study. For example,

Approximate to:dcity order was E!.. t~tr..9.§',

The first three responded to 1.7 uM (0.1 ppm) and

the other to 10.2 uM (0.6 ppm) by reduction in growth. Skaar ~t

by 0.5 ppm of the metal. The species showed slight growth

reduction at 1.0 ppm of the total Ni.

Hutchinson and Stokes (1975) reported the sensitivity of

~£~!1~Q~2ffi!:!2 ~£'l:!!!!iQ~t9: to Ni among many species of Algae studied.

~b!Q~~!!§ ~l:!!g§r..i§ was the most tolerant one observed in the

study. When the growth rat.e of §~ ~£l:!miQ~t~ was affected

significantly by 0.05 ppm of the metal, more or less equal

response was obtained in case of gb!Q~~!!~ ~l:!!g~r..i§ only by 0.3

ppm. §~ ~£l:!min§t~ got inhibited by 94% in growth rate by 0.1

ppm and gb!Q~~!!~ ~l:!!g§r..i2 showed a reduction in growth rate by

54% in presence of 0.7 ppm metal.

51

Hollibaugh ~t §l· (1980) r~ported that 1000 nM (0.06 ppm)

They observed the

toxicity of Hg to be many magnitude higher than Nito the same

species. Inhibitory effects of Ni o~ chlorophyll was also

demonstrated in the study. Whitton and Shehata (1982) observed

different sensitivities to different stock of blue green alga

~n~£~§ti§ OiQ~l~o§·

ppm of Ni strongly.

Growth of wild type was inhibited by 0.16

Stokes ~t §l. (1973) had also noted similar

observations in case of §£~n~Q~§m~§ ~££~min~~~ from different

stocks. Inhibition of §!l~n§§~L~m £~QLi£QLO~~~m by 0.40 ppm Ni

was demonstrated by Michnowicz aMdWeaks (1984) at different pH.

The optimum pH showed lesser toxicity than lower or higher pH

conditions. Study conducted by Spencer and Nichols (1983) show

an inverse relation between 14th day cell number and free

Ni+2 concentrations.

by 10.2 uM Ni(T) (0.6 ppm) in the ~bsence of chelators. Sparling

(1968) conducted study on four species of blue green alga, viz.

and observed significant reduction at only

34 to 170 uN of the metal (2.0 - 10.0 ppm). The higher value

necessary for significant inhibition of green algae seen in this

case is due to high EDTA present in the medium. Inhibition of

the growth of QblQL~ll~ ~p. with Ni was reported by Upitis ~~ ~i.

(1974). They observed insignificant reduction of growth at 0.1

,

52

ppm of the metal. At 1.0 ppm the growth was reduced by ~ 4.4% of

the control and 2.0 ppm by ~ 53%. Ishizaka ~t ~l·

compared Ni toxicity to ~blQL~ll~ ~Y19~Li§. It was reported a

higher toxicity of the metal than Pb,

lesser than Hg,. Ag, Cu and Cd.

Co,

In experimental stream conditions exposed to natural

physico-chemical conditions Patrick ~t· ~l. (1975)' demonstrated

the alteration in species composition due to Ni e:t;posure. .An

increase of the green and blue green algae' and a decrease in

species diversity was induced in the community by 0.062 to 1.0

pprl' Ni.

Significant decrease in cell size, increase in the surface

area of cell thylokoid, and reductions in the volume of intra

thylakoid spaces, coalescence of cellular lipid etc. was induced

by Ni exposure to El~£tQn~m~ bQL~~nYm (Rachlin it ~l. 1'382) upon

exposure to 100 ppm for 4 hrs. Whitton and Shehatha (1'382) also

reported morphological modications by Ni in an~£~§ii§ ni~Yl~n§.

FOY'mat i on of filaments upon exposure to partially inhibitory

concentration was observed. (1982) reported

formation of polyphosphate bodies in case of El~£tQn~m~ bQL~~nYm

(cyanophycat?-) . Flavin and Slaughter (1974) reported inhibition

of flagt?-llay movement in ~b~l~n~~QmQn~§ L~inb~L~ii with 0.18 mM

nickel acetate, with 0.6 mM hindrance of flagellar detachment and

with 0.3 mM, inhibition of flagellar regeneration. A loss of

53

coordination between flagella was observed by Bean and Harris

( 1'377) resulting in abnormal swimming behaviour

On a species of marine bacteria ~~ih~Qb~~tQ~ m~~inY§

alteration in cellular structure, morphology and growth pattern

was observed upon Ni exposure (Cobet ~t ~l. 1970).

Qhr..QmtYm

A number of studies on to~/;icity of C:r have been conducted on

different organisms (e.g. Eisler and Henneckey, 1977; Fales,

1978; Frank and Robertson, 1'37'3 ; Pickering, 1980; 50ni and

Abbasi, 1981;Ramusino ~t ~!.i 1981; Oshida and Ward, 1982; Abbasi

and Soni, 1'383; Pagano ~:£ ~!. 1983; Bianchi and Levis, 1984;

Bookhout ~t ~!. 1984). But most of the studies were concerned

with invertebrates, fishes and higher organisms. Aquatic

toxicological studies of this metal especially with emphasis to

phytoplankton groups both under laboratory conditions and in §it~

.conditions are very scarce.

Among the two main stable valence states of Cr (Cr-3 and

Cr-S ) the higher valency state is found to be more toxic (Mertz,

1959; Towill ~t E!.. 1978) • Cr -6 is relatively more stable in

water (Cutshall ~:£ §!. 1955; Fukai, 1957). The species is more

toxic because mainly of its higher oxidation potential and ease

in penetrating biological membrane (Rollinson, 1955; Mertz, 1969)

and high stability in water (Bookhout ~t ~l. 1984). On the other

hand, trivalent Cr is mainly found bound with particulate matter

and the reduction of higher valent form to the lower valent stage

is helped by the particulate matters (Curl lE.t g1.. 1965;NF~C,

1974; Pfeiffer lE.t g1.. 1-'380, 1 '382) • On uptake into biological

system it 'is reported that the hexavalent form gets reduced to

the trivalent form (Mertz, 1969; NPC, 1'374). Bringmann and Kuhn

(1959) reported to toxic threshold concentration of 5 ppm

trivalent Cr to §£~O~~~§mY§ SPa and reduction in photosynthesis

rate of ~~£rQ£~§ii§ Q~riilE.r~ with 1.0 ppm of hex~valent Cr.

Wiuro-Andersen (1974) conducted a study on the effect of Cr on

the photosynthesis and growth of diatoms and green algae. The

species engaged in the study were gb1.Qr~1.1.~ Q~r~oQi~Q§g and a

In case of the diatom 0.15 ppm metal

inhibited the growth significantly. The cell number was found to

have important role in case of Cr toxicity also. 1.0 ppm Cr

reduced the photo-synthetic activity of the diatom by 70%. Ten

times more Cr was necessary for a more or less equal effect on

the green algal species. Wium-Anderson (1'374) concluded th~t Cr

is less toxic than Cu to both of the species. YOt1gue lE.i ~l..

(1979) explored the joint effect of temperature and Cr on ~ygl.~n~

Increase in metal concentration reduced the survival

rate of the organism. Concentration upto 1.0 ppm showed no

significant effect on the survival of the species. Yongue ~t

~l.. (1979) state that heat treatment potentiates the toxicity of

Cr. Nollendorf lE.i ~l.. (1'372) observed that Cr at concentrations

of 9.5 uM/I (0.5 ppm) inhibited the growth of gbl.grlE.1.1.9. Spa

55

Inhibition of §~!~n§~~riYm £~QLi£QLOYtYm by Cr concentration of

9.5 u mole (0.5 ppm) was reported by Garton (1973)."

Patrick ~~ ~!. (1975) observed inhibition of diatom growth

in a mixed population by 1.9 u mole (0.1 ppm) Cr and a complete

replacement of diatom population by blue green algae at 7.6 u

mole (0.4 ppm). In an artificial stream microcosm experiment

Patrick (1978) reported the gradual change of dominance in the

community from diatom to blue green alga, with an increase in Cr

concentration from 49.5 ug/l to 376.5 - 405 ug/l.

(1983) conducted C:r toxicity study on "marine

phytoplankton assemblages and Ib~!i§§iQ§iL~ Q§~Y~Qn~n~.

natural

In high

salinity

observed.

(37.5%) no effect of Cr upto 1.9 uM (0.1 ppm) was

Only at a concentration of 19 uM/1 (1.0 ppm) a lag in

growth was observed. §t~!~tQn~m~ £Q§t~tYm was eliminated at this

low salinity experiments 1.9 uM (0.1 ppm) Cr concentration. At

decreased the growth and 0.19 uM/l (0.01 ppm) affected an

apparent lag in growth bf phytoplankton. The main species

inhibited at the concentration was §YLiL~!!~ Q~~t~L Q~!QmYi~

£Qnf~L~~£i~" and ~~£!Qt~!!~ Spa At lower salinity (0.03%) 0.19

uM/1 (0.01 ppm) Cr inhibited growth of Ib~!~§§iQ§iL~ Q§!Y~Qn~n~

and at 1.9 u mole/l (0.1 ppm) the growth was severely inhibited.

Hig~er salinity showed an ameliorating effect on toxicity.

show

bioconcentration

high

and

capa~ity of heavy

biotransfer. They

56

metal uptake,

show high

bioconcentration capacity in relation to the biomass (Ribeyre and

Boudow, 1982) mainly due to the high surface area to biomass

ratio. Micro algal cells have diameter in terms of micrones and

specific geometric surface area in ordei of m2/g fresh weight

(starry and Kratzer, 1984). Davies (1978) suggested that the

cellular surface of phytoplankton consisted of a mosaic of

cationic and anionic exchange sites. The net charge on .the

surface layer determine to a great extend the accumulation

capacity of the alga to heavy metals. So the pH of the ~edium

and other parameters show an important effect on determining the

accumulation capacity. Babich and stotzky (1980) states that the

increased pH of the medium decreases the competition between the

cations and protons,

~ptake of toxic metals.

thus e~hancing the binding capacity and

Micro organism possess two processes in heavy metals uptake,

~he first involving non-specific binding of metals to the cell

surface, slime,jayers and extra cellular metrices and the second

metabolism dependent intra-cellular uptake mechanisms (Bollag ~nd

DU5zota, 1984). The uptake of heavy metals at the initial stage

is relat~d to simple ion exchange process (Jennett ~i ~!. 1984)

and is a passive process (Glooshenko, 1969) where cations replace

57

those on the cell surface. The exchange const~nt KCH- M ) of the

respective cationic species to the binding sites in respective

conditions determine the process. Starry and Kratzer 0'384)

state that the algal cell wall behave like a weekly acidic

cation exchanger with various cell wall legands of different

capacity. The increase in pH leads to the exchange of protons to

cations present in the medium. Tht?y observed that the algal cell

walls in natural conditions predominantly are in Ca and Mg forms

with Na and K ions occupying only 1-3% of capacity. Bryan

(,1971) assumed the alginates (Uronic aci,d polymers) present i t1

the cell wall and intracellular spaces of b~miQ~~i~ ~igit~t~

work as ion exchange materials. E~/;tracted algi nates showed the

following order of affinity as shown by the whole cells.

Pb > Cu > Cd > Ba > Sr > Ca > Co > Ni > Zn > Mn / Mg

After the binding of the metal cations to the cell surface the

transport is affected by either passive or active process

depending mainly on configurational similarities of the element

with some of the nutrient compounds resulting possibly in a

competitive interaction (Hart ~t ~l. 1979; Nielsen, 1980).

Algae show lesser accumulation capacity to anionic species

(Starry ~i' ~!.. 1984) and a de~rease in the capacity with an

increase of the equilibrium pH and concentration of the metal

spec i es. The observation is significant in case of uptake of

58

cE-rtain hE-avy filE-tals which can form anic..nic radicals in aquous

medium likE- er, W, Mb, E-tc.

GenE-rally all mE-tals arE- concE-ntrated by algaE- to a cE-rtain

e:.o:tent, thE- dE-grE-e varying with metals and algae (JE-nnE-t ~t §l.

1'382) • JennE-t ~t §!. (1'382) observed concE-ntration factors (CF =

C/C:' , whE-rE- C = concE-ntration in the organism, C'= conc. in thE-

mE-dium) widely varying with diffE-rent mE-tals. For some mE-tals the

CF was found to vary by a magnitudE- of thousands to diffE-rent

speci es (e. g. , 1982).

Variation in concE-ntration factor with gE-nus and mE-tal was also

demonstratE-d by TralloppE- and Evans (1976l. A relation between

Cd uptake and Ca and Mg was dE-monstratE-d and the influence of the

hardness of the medium was rE-portE-d in case of ~it~ll~ Spa by

Kinade and Erdman (1975).

(1984) observed a direct relation of uptake of

Ag- and their ionic radii. They found· an exception in thE- case

of only Cr 3 + possibly due to the fact th~t actual charge of Cr

complexes in aquE-ous medium is {3 and the ionic radii are

significantly largE-r than that for single Cr+ 3 ions. pH

dependence of accumUlation of metal (Zn and Cd) was rE-portE-d b~

Starry ~t §!. (1'383) • ThE-Y found that the logarithm of thE-

conCE-ntration (F) is a linE-ar function of thE- pH valUE-. Starry

~i ~l· (1983alalso rE-portE-d thE- pH indepE-ndent naturE- of mercury

59

They ,supposed that this

different effect of pH on metal uptake is due to the fact that at

the conditions of the study (pH 4-9) most of the Zn and Cd are in

ionic form and Hg in Hg(OH)2 or HgCl 2 form. The algal cell

density found to have no influence on the metal .=oncentration

factor in their study. Starry ~i el· U'383b)

reported a concentration factor of 4500 at 1.0 mg/l level.

Sakaguchi (1979) discussed about an indirect correlation

between algal cell density and concentration f act or. Stokes

0'375) reported a concentration factc.r of 1 y; 104 for Ni -by

Hardy ~i £!... (1'384) reporte·d 1680 as the

Geisweid and Urbach (1983) discussed about the variability of

concentration factor .for Cd with the cell volume of ~blQ~~!..!..~.

An increase in Cr-a uptake with pH increase from 6 to 7 was noted

by Starry ~i £!... (1983a).

Davies (1973) studied the dynamics of Zn uptake by

An increase in uptake of the metal

linearily with the square root of time in the initial phase of

uptake process was observed. Davies (1973) demonstrated that the

metal absorption by Eb~~Q~~£i~lYm follows Langmuir isotherm and

concludes that the metal uptake in the initial stages is a

passive process and dependent on the intra-cellular protein

content. De Filippis and Pallaghy (1976) reported that for

so

~b!QCQ!l~ yy!g~Ci§ maximum Zn sorption was within 1--2 hI's.

Simi lar was the observation by Saka-guchi ~i ~l· ( 1 '37"3) on the

cumulation of Cd by same species. In this case the maximum metal

uptake occurred within 30 mins. Gipps and Coller (1980) observed

on the same species maximum uptake within 8 mins bf exposure.

They distinguished a 'rapid uptake' and 'slow uptake' phase.

They propose that only at later stages the passive absorption or

diffusion through the cell membrane or active transport process

become important.

Geisweid and Urbach (1'38"3) reported a fast and slow phase of

sorption of Cd by Gb!Q[Q!l~ yy!g~Ci§L ~nti§i[Q~Q§mY§ ~[~Ynii and

A decreasing effect of cell volume and an

increase of the uptake was observed. Geisweid and Urbach (1983)

demonstrated that Cd sorption at equilibrium can be described by

Freu~dlich sorption isotherm. More or less <::.--qual or more

sorption was seen in case of dead cells. They propose that metal

sorption can better be demonstrated by cOE'fficients of

Freuendlich or Langmuer isotherm rather than the accumulation

ratio which are found to vary widely with cell volumes and free

metal concentration. ImportancE' of surface sorption rather than

the energy requiring process for metal uptake was reported by

Gadd and Griffith (1978).

Many studies reported that thE' metal sorption by dead cells

is mOl'e or lE'sS higher than the living cells indicating thE'

61

importance of passive sorption process (Bentley-Mowat and Reid;

1977; Button and Hostetter, 1977; Kurek ~t ~l. 1982; Bollag and

Duszota, 1984) ..

Rebhum and Ben-Amotz (1984) observed that Cd absorption by

algae is dependent on the itlitial concentration of Cd and it

increases with the metal concentration. They found that the

following mathematical relation is followed by Cd distribution

Cd content in algae, pg/cell; C = residual Cd concentration in

the mediuril, K == a constant; n :::: e:.-:ponent i al constant) ..

This expression was found applicable at concentration ranges of

1-10 ppm Cd. In their system they found the relation to be Y =

1 • 00 :.-; C 1 • 7 • A positive value for the exponential constant

indicated the increase in uptake with the concentration. Hawkins

(marine phytoplankton species) observed faster uptake phase (5-10

mts) followed by a slower uptake phase upto 2 hrs and then after

no further uptake upto 24 hrs. The rapid uptake they state is

due to the adsorption of the metals to the 'apparent free space'

or apoplast and it can be rt:<moved by EDTA washing. It was

50% in §~ ~~£ill~~i§ was firmly bound copper which

cannot be removed by EDTA treatment. Hawkins and Griffith (1982)

concluded that some algae like Q~ ig~tiQlg£i~ have effective

preventive mechanisms against copper uptake. Fiom their study

sensitivity of the algae and the capacity to restrict copper

uptake was found to be related.

A saturation effect of metal accumulation was noted in case

(Dongmann and

Nurnberg, 1'382). With increase in concentration of the metal in

the medium in the range of 0.4 to 25 uM, a decreasing effect on

accumulation factor was observed. They observed different

saturation constants for Cd accumulation by the.addition of metal

at different times (accumulation data was fitted to a Langmuir /

adsorption siotherm to determine the saturation constants. K =

K __ >< C/(K....:: + C) wher€:' K and K ........ '" = the molality and the rlld.:,;imum

flrHolality of t,he metal in algal biomass. Kc = saturation constant

= the metal concentration at which sorption corresponds to half

the maxim~m sorption). 10 uM/1 EDTA was found to have no effect .

on the metal accurnulation in the case of Ih~l.~~~iQ~ir:..~ r:..Q:t!::!l.~. ,

Dongman and Nurnberg (1'382) conclude that Ni is bound strongly to

the species by a factor of 100; and desorption probability of Cd

is in two order higher than Ni.

Les and Walker (1984) studied the short term binding

capacity of a sheeth producing blue green alga ~hr:..QQ~Q~~Y~ Q~r:..i§.

They detected that: 90% of Cu and Zn was bound within 1 min and

most of each metal within 10 mins. Washing with dilute EDTA

solution removed '38-100% of the' metal sorbed within a few

63

minutes. Among the three metals the binding affinity was found

in the order Cd > Cu > Zn. Upto 4.0 ppm Cd, S.O ppm Cu and 20.0

ppm Zn the Fruendli~h isotherm was found to be linear.

Concentr-ation factors determined showed a saturation effect with

metal concentr-ation incr-ease. pH influence on binding capacity

was comparatively insignificant. Concentr-ation factor-s for- the

species was Cd ~ 3200; Cu ~ 3800; Zn ~ 4200.

Using radiometric methods Starr-y ~i ~l. (1':::J83b) studied

accumulation of 14 elements including Cd, C:u, Zn, Co etc. by

They also summarized

data of 42 species of 30 elements on the accumulation by ~~

C' ;:).

in~gLti~ from studies conducted by their- group. In rllost of the

cases the equilibrium F values (cumulation factor) wer-e attained

within 30-60 mins. HgCl 2 and phosphate. Alkaline

and alkaline earth metals got cumulated at the same sites 6f

algal cells with approximately same capacity. A competitive

inter-action b~tween the metals was observed. The cumulations of

metals like Cu and Cd and other hydr-olyzable cations, with a rise

in pH, a linear- increase was observed. The maximum of the

process reached at pH > 9. Algal cells are reported to have

capacity to accumulate M2+ and M(OH)+ forms of metals. The

64

IOrder. of cumulation depend on equilibrium pH. At pH '" 7 the

order was as follows:

Cr .... 3 Fe3 -+ Ce::3+ Cu2 -+ ) Ag >- Pb TI+1 .> M2+j Co2 .... >- Zn; Ba :::- Cd.

They observed that the cumulation of metal can be described

quantitatively by using the affinity constants (K ... ) : maximum

cumulation factor o~ maximum capacity of the algal

cumulate metal species (CAM).

cells to

Oi f fe-rent i al accumulation of heavy metals by epiphytic

During the first 4 hI's, uptake increased in case of

metals Cu (0.01 - 0.04 ppm); Pb (0.01 - 0.06 ppm); Zn (0.015

0.05 ppm);

0.06 pprrl);

Ni (0.02 0.04 ppm);

Co (0.055 - 0.075 ppm);

Cr (0.04-0.06 ppm); Cd (0.04

Mn (0~05 -0.10 ppm) and Fe

(0.05 - 0.1 ppm (in brackets the concentration of the metal

appl i ed) • In this study three course of metal uptake was

distinguished (1) in which rapid uptake within 1 hI' followed by

slow uptake and (2) 2 hI's rapid upt~ke then slow uptake and (3)

continuous uptake during the entire 4 hI's of exposure (Ni, Cr, Fe

atld Mtl). In case of Ct the uptake was found to be very less.

Only 49.5% of the metal was removed during the whole stDdy.

Hart and Scaife (1977) studied the bioaccumulation of Cd by

£b19Lgll~ Q~L~~9iQ9§~·

of Cd to ~blQLgll~ Spa

The pH increase, decreases the toxicity

A pH dependent uptake variation with

65

lesser uptake at higher pH was also noted for the algae. A two

fold increase with 1 unit decrease i~ pH was observed. They

found a directly proportional increase of uptake with external

concentration. Mn concentration alters the Cd uptake. Above 0.2

ppm of Mn no uptake of Cd was seen. Neithe~ Ca nor Mg altered

the uptake. Similar was the case with Zn, Co, Cu and Mb. Mn and

Fe appeared to function as competitive inhibitor of Cd (Hart ~~

197'3) . A possible correlation between uptake and toxicity

.was observed. Factors which enhance uptake result increased

toxic actions also. Inhibitions due to Cd on CO 2 fixations and

O2 release was comparatively small when compared to the uptake.

The possibility of some specific metal binding proteins were

suspected for this reduced toxicity even , after high uptake. A

concentration factor around 5000 was observed during the study.

A higher uptake of copper by the non-tolerant variety was

observed by Jensen ~i ~l. (1'376). A positive correlation of

temperature and metal content was observed in ~blQL~ll~ and

Th~ maximum uptake rates was in the

range of 29.5 (optimum

lRibeyre and Boudow (1982) also reported the temperature-metal

uptake relationship. The tolerent strains of algae are found to

accurflulate more Cu than the others O:::lotz, 1981). Stokes (1975)

discussed about the variation in Cu uptake by tolerant and non

t 01 er ant al gae. Shehata and Whitton (1982) observed that uptake

66

by the tolerant ~O~£y§!i§ oi~~1~o§ was more or similar to that of

non-tolerant strains. They ascribe the tolerance observed to the

capability of the algae to sequester the metal in some for rl",s

which do dot interfere with the metabolism. In case of

of eu in sensitive strains.

Skaar ~:t. 21· (1974) states that the uptake of Ni depends

strongly on the metabolic state of Eb~~Q~2£!y1~m :t.~i£Q~o~:t.~m. Ni

~ptake was found to be related to"the pho~phate availability.

Phosphate starved algae showed less Ni binding. capacity. During

the first 10 hrs the uptake rate of the metal was the highest. it

is supposed that phosphate play important role in Ni binding

system of the algae. The Ni uptake was significant even when Ni

applied was as low as 0.0005 ppm

responded to only 0.5 ppm.

(NiZ+) while the growth

When an organi sm is e:/;posed to more than one poll utant

si rfJul t aneousl y, it results in modifications of the toxicity

expression of each of the toxicant, and can be (a) an altered

effect unpredictable from the components toxicity or (b) at1

effect different from that of the constituents in the mixture.

Sprague (1970) categorized the multiple toxicity into: (a) more

than additive (potentiation); (b) joint action (synergism); (c)

67

additive; (d) less than additive, and (e) antagonistic effects,

after Gaddum (1948). Finney (1971) and Andersen and d'Apollonia

(1978) also discussed about different types of multiple toxicity

like- (a) strict addition - the mixture toxicity is similar to

constituent toxicity and these effects combine additively in the

mixture, strict addition also involves supra additive, a form of

synergism and infra-additive - a form of antagonism, (b) resp()t1se

addition - the toxicants act on different systems but produce

common response, (c) sensitisation and potentiation - in which a

non toxic pollutant promoting either the binding or toxic action

of another on~ and (d) permissive synergism where pollutants

interacts and produce an effect different

toxi.: i ty.

from individual

Synergi st i c interaction between different metal combination

on various species of algae were reported by Break ~i ~!. ( 1976)

Christensen ~t ~!. ( 1'379)

I!:l~lS!§.§.iQ§.ir.~

(§~l~t222tr..~!!!

~!. (1974) (~~!~t2~§.tr..~!!! ~~~r.i~Qr.t2~tYm Cu and Cd), Say and Whitton

,(1977) Cd and Zn), Stokes <: 1975)

eu and Ni). Antagonism was observed by Bree-k

<: 1976)

atld Morel ( 1983) Cu and Cd) ,

Christensen ~i ~!.

C:ii pps arId Bi I'" 0 <: 1'378)

68

Cd and Mn), Deviprasad and Deviprasad ( 1982)

Cd and Pb:> • Shehata and

Whitton (1982) found additive action on An~£~§ti§ Qi~~l~U§ with

Cu and Zn; Ni and Zn; and Pb and Zn combination.

The metal-metal interactions leading to different types of

toxic effects as discussed above are dependent on the

relative concentration of the toxicants and (b) the sequence of

exposure to the toxicants (Babich and Stotzky, 1980:> • In the

varied manifestations of the combined toxicity, the interactions . . occurif in environmental phase,

( kt.net i c phase· and lor dynami c phase

(Anderson and d'Apollonia, 1'378). In the environmental phase the

interaction alters the bio-availability of the toxicants, the

kinetic phase includes interactions commencing with the binding

of the toxicant to the target tissue an~ dynamic phase, the

interactions which determine~ the availability of the toxicant to f

different body compartments.

Nielsen (1980) distinguish~s in between competitive and non-

competitive interactions which occur in between Ni and different

metals like C:a, Cr, I, Fe, Mg, Mn, Zn, Mb, P, K, Na, etc.

The competitive interactions are affected due to the physico-

chemi cal similarity in between the competing elements leading

into isomorphous exchange at the functional si tes. f\nd non-

competitive interactions are affected if the djfficiency of one

69

element alters the biological .functionsof the other. Medon et

0'384) discussed about the possibility of competitive

interactions between Ca and Pb to effect a decrease itl the

toxicity of Pb to §Qb~~~Q1ily§ Q~1~U§ (Bac t er i um) .

interference is the possible mechanism of the enhanced eu

toxicity observed by Sunda and Huntsman (1983) due to low Mg

concentration in cultures. Th~ interference of silicic acid

uptake by eu

manifestation of the interference of5ilicio:- acid metabolic ,cycle

rather than competition for binding sites (Rueter et ala 1981) =

The situation of complementary accumulation and antagonism

observed in case of Se and Hg by Leonzio ~t ~l.. (1'384) is also an

example of competitive interaction.