Interactions of lead(II) with natural river water. Part II: particulate matter

)) . ,~

>

~.,'~

the Science of the Total Environment

.u'"

ELSEVIER

The Science of the Total Environment 151 (1994) 101 112

Interactions of lead(II) with natural river water. Part II: particulate matter C.M.S. Botelho a*, R.A.R. Boaveotura a, M.L.S. Sirn6es Gon~alves b, L. Sigg c aDepartamento de Engenharia Quimica, Faculdade de Engenharia, Rua dos Bragas, 4099 Porto Codex, Portugal bCentro de Qulmica Estrutural, Cornplexo I, Instituto Superior T&nico, At/. Ro~sco Pais, 1096 Lisboa Codex, Portugal Cinstitute for Water Resources and Water Pollution Control (EAWAG), Swiss Federal Institute of Technology, CH-8600 Diibendorf, Switzerland Received 18 January 1993; accepted 10 May 1993

Abstract

The distribution of lead(II) between particulate and dissolved phase in the River Este (Portugal) was studied by field measurements and compared with calculated simulations, using parameters obtained by adsorption experiments with the suspended particles and with dissolved organics. Experiments in which Pb(II) solutions are titrated with a suspension of particles and analysed by DPASV are interpreted in terms of binding capacities and conditional stability constants of Pb(II) with surface sites. Desorption of Zn(II) ions observed during experiments with Pb(II) made it possible to determine Zn(II) complexometric surface parameters. From speciation calculations, using experimentally determined complexometric parameters, we concluded that dissolved organics and particle surfaces control lead distribution. For Zn(II) the free species exists in higher concentration and metal distribution is not very sensitive to changes in the composition of the solution.

Key words." Surface complexation; Heavy metals; Suspen'ded particles; River pollution

I. Introduction Industrial activity provides enhanced fluxes of several trace metals to natural waters occurring in dissolved, colloid or particulate forms with different biogeochemical properties that determine the influence on their mobility and bioavailability (F6rstner and Wittman, 1981; Tessier et al., 1989). An understanding of the processes and factors involved in the mobility of trace metals in the aquatic environment is necessary in order to make

* Corresponding author.

management decisions, e.g. determination of waste levels and disposal of dredged material. A large fraction of trace metals such as lead(II) and zinc(II) introduced into the aquatic environment is usually associated with particulate matter that settles down onto the sediments, although little is known about the processes of scavenging of trace metals by particulate matter, their settling to the sediments and release to overlying waters. Voltammetric methods such as differential pulse anodic stripping voltammetry (DPASV) are the appropriate techniques for determining the extent of interaction between metal ions and particles, because in natural water conditions the

0048-9697/94/$07.00 © 1994 Elsevier Science BV. All rights reserved. SSDI 0048-9697(94)3950-7

102

CM.S. Botelho et al. / The Science of the Total Enviornment 151 (1994) 101-112

measurement can be made without previous filtration or centrifugation that cause laboratory artifacts (Gon~alves et al., 1985). In Part II of this work a study of surface complexes of lead (II) with natural particles of three different sites along a polluted river was carried out in order to complement the work performed with soluble organics in Part I. To provide a direct measure of the metal adsorption onto the voltammetric cell components a n d / o r precipitation we inverted the titration method used in Part I. This procedure has also been applied by Miiller and Sigg (1990). Despite the advantages of this method: assessment of the adsorption equilibria at low concentration of metal ion, as well as minimization of the effects of polymeric species in solution and on the surface, complexes with lower stability are quantified only in the beginning of the titration when the ratio metal/ligand is high and the experimental errors are easily propagated. The two titration methods with the metal ion and with particles are complementary and it is best to use both of them for the same sample. Zinc desorption was also studied because natural particles have zinc adsorbed and the influence of organics adsorbed on particles was also considered. Speciation in the natural waters was analysed, and the distribution coefficient experimentally determined in the field and calculated from the stability constants was compared. 2.

Theory

The concentration of total metal ions adsorbed on particle surfaces (E[-=SJLM]) can be obtained from the mass balance: CMT = C M + Cce I =

E[=-SJLM] + (1 + Kc)[Mso 1] (3)

where C M is the metal concentration that is available for complexation in solution and with particle surfaces, and CceI is the amount adsorbed on the cell or precipitated. Dissolved organics can exist in solution by desorption from surfaces. [Msot] represents all species in solution, free metal and labile metal complexes: [Mso,] = [M](1 + EgH[OH- ] n + ]~3[CO2-] n +EJK[YL])

(4)

For surface groups the mass balance is: [--SL]T = E[=SJL] + E[-SJLM] = Cc P

(5)

where Cc is the surface binding capacity expressed in mol/kg of solid material and P is the concentration of suspended solids (kg/l). For labile complexes in solution EJK[JL] in Eq. 4 can be obtained from the shift on the peak potential using the DeFord-Hume method (DeFord and Hume, 1951). If all the surface sites are assumed to be equivalent, all the JK M values are identical and JK M = K M. If this is not possible, a mean value for the conditional stability constant can be defined as:

2.1. Adsorption of metal ions on solid surfaces Simple model assuming discrete distribution of heterogeneity.The adsorption process is inter-

KM = EJKM[---SJL]

preted in terms of surface complex formation of several monodentate surface complexes:

The [-SL] values are obtained from Eq. 5: [-SL] = Cc P - [-=SLM] and after rearranging Eq. 6 we can obtain:

---SJL + M ~ - S J L M

JK M

[=-SLM] [M][-=SL]

(6)

(1)

JKMiS the conditional stability constant at constant pH, for the complexometric reaction of metal ion with the j surface site and can be defined as: [=SJLM] JK M = [M][_SJL]

Z[=SJL]

(2)

{~SLM} [M] = K~Cc - KM{=SLM}

(7)

where {=-SLM} is the concentration of adsorbed metal expressed in mol/kg of solid material: {-=SLM} = [=SLM]/P. Eq. 7 is an expression of the Scatchard method (Buffle, 1988). Therefore, from the plot of {-SLM}/[M] vs. {-=SLM} a

C.M.S. Botelho et al. / The Science of the Total Enviomment 151 (1994) 101-112

straight line is obtained and the values of K M and Cc can be obtained. The determination of the complexometric parameters can also be carried out from the plot of [M]/{SLM} vs. [M] (Berg and Kramer, 1979; Rfizic, 1982; Lee,. 1983), where slope = 1/Cc and intercept = 1/CcK M. Because slopes are always determined with higher accuracy than intercepts, the Scatchard plot is more suitable to determine stability constants than the plot first proposed by Berg and Kramer (1979), and the latter is more suitable to determine complexometric capacities. Eq. 7 can be transformed into: {-=SLM} = Cc

KM[M] 1 + KM[M]

(8)

which is an expression of Langmuir isotherm, valid only for equivalent surface sites with no interaction between neighbouring sites and for the formation of a monolayer of solute on the solid surface. If there exists two different surface groups (with surface binding capacities l Cc and :Cc, respectively), concentrations of bound and free metal are related to each other by (Buffle, 1988):

{~SLM}=ICc 1KM[M] +2Cc 2KM[M] 1 +1KM[M] I+2KM[M] (9) where 1K M and 2K.u are the conditional stability constants for the complex formation of metal M with the dominant sites 1 and 2. When this occurs the plots {-=SLM}/[M] vs. {-=SLM} or [M]/{=-SLM} vs. [M] will be curves and complexometric parameters can be obtained from the analytical expressions for slopes and intercepts of the straight lines relative to centers 1 and 2, separately (Rfizic, 1982; Buffle 1988). Heterogeneous model that assumes a continuous distribution function of sites. As has been described in Part I the differential equilibrium parameter is a more precise measurement for complex stabilities than the average equilibrium coefficient. In solution, a large number of different complexing sites on natural particle surfaces has to be as-

103

sumed, each one interacting with the metal ion in a different way. During titration with particles the stability of surface complexes must increase continuously. With the increase of the active sites number, for a constant total metal concentration, metal ions leave sites with minor affinity and bind to those of higher stability.

2.2. Organics desorption from surfaces The adsorption equilibrium of organics on particle surfaces can be expressed by: ~S + L ~ - S L

(10)

and an equilibrium adsorption constant can be defined as: [---SL] K~ = [-=S][L]

(11)

according to the Langmuir isotherm. The addition of =SL and =-S groups to the solution is due to the titration with particles. Desorption of species L occurs until the concentration of L in solution is so high that the subsequent desorption will not be significant and the desorption isotherm attains a flat region. If equilibrium concentration of desorbed L and particles surface complexometric capacity are known, Ks can be calculated, assuming that it is a constant for all particle concentrations. Calculations can be undertaken assuming that for all suspension concentrations the initial (before complexation by metal ions) ratio: [-SL]0 -

-

[---S]0

=R

(12)

is a constant. For each particle concentration, mass balances can be written as: [---SL]0 + [~-S]0 = CcP

(13)

once equilibrium concentrations of desorbed L (CL) can be determined. Ks can now be defined

as" Ks =

[~-SL]0 - C L CL([~--S] 0 -}- e L )

(14)

The value of C L can be found from voltammetric titrations with the metal ion of solutions of organ-

104

CM.S. Botelhoet al./ The Scienceof the TotalEnriornment 151 (1994)101-112

ics desorbed from the particles after previous filtration, as described in Part I for the dissolved organics in the river water.

3. Experimental 3.1. River system The sampling sites have been described in Part I of this work (Botelho et al., 1994). 3.2. Sampling Suspended particles were obtained by batch filtration (0.45 /~m) of water samples. Filtration was carried out within a few hours of the sampiing after return to the laboratory. These particles were then freeze-dried and kept for use in laboratory experiments and total metal concentrations were determined. All glassware coming into contact with samples was checked for contamination; bottles, flasks, cells and filtration device were cleaned with HNO 3 1:5. 3.3. Analytical techniques Organic carbon in particles was determined on a concentrate suspension of dried particles in deionised water. Suspensions were homogenized with an ultrasonic disintegrator and analysed for TOC by catalytic conversion into C O 2 (900°C) and the gas was measured by IR spectrometry on an IONIC-1258 T C / T O C analyser. Inorganic carbon was removed beforehand by acidification and bubbling with N 2 gas. Total phosphorus in the particles was determined spectrophotometrically by the ascorbic acid/molibdate method after digestion (APHA et al., 1985). To determine total metal concentrations, particles on filters were digested with concentrated nitric acid and analysed by atomic adsorption spectrometry for Fe, Mn, Cu and Zn. The solutions of digested particles were analysed for Pb by DPASV, using standard additions method and a Metrohm 506 polarograph coupled to a Metrohm 663VA polarographic device. 3.4. Experimental procedure Pb(I-I) adsorption.In order to study adsorption of Pb(II) on natural particle surfaces, metal ion

was allowed to react with a suspension of the filtered particles; conditional stability constants and binding capacities being determined. Conditions were chosen to be close to natural ones. The procedure used according to Miiller and Sigg (1990) is as follows. For particles L and F, solutions of lead(II) (2.2 X 10 .7 M in 0.01 M KNO 3, 0.001 M NaHCO 3 and buffered with N2/0.006% CO 2) were titrated with increments of a suspension of particles (1-40 m g / l in the final suspension) at constant pH (pH--7.4 and pH = 7.7, respectively). Owing to the higher pH value of sample B (pH = 8.2) a greater adsorption of metal ion on the cell device and on the particle surfaces was anticipated and therefore a higher concentration of Pb(II) was used. A solution of 4.4 x 10 .7 M in Pb(II) and 0.01 M in KNO3, buffered with boric acid 9 × 10 .4 M and sodium borate 8 × 10 .5 M, was titrated with increments of particles (1-40 mg/1 in the final suspension) at constant pH (pH = 8.2). Dissolved lead concentration (free metal and labile complexes) were determined, 20-30 rain after each addition. A Metrohm 647VA polarographer coupled with a Metrohm 646VA processor with the hanging mercury electrode (HMDE), using DPASV was used. Oxygen was first purged from 20 ml samples for 15 min using a mixture of N 2 with 0.006% CO 2 for samples L and F and pure N 2 for sample B. The deposition time on the HMDE was 80 s at - 1100 mV for zinc and - 800 mV for lead. The scan rate was 10 m V / s and the pulse amplitude 50 mV. A glassy carbon electrode was used as a counter electrode, the Ag/AgC1/KC1 reference was a PTFE-capillary with diaphragm and KNO 3 filling solution. Adsorbed metal ion is discriminated from the measurement due to the size of the particles and to the significantly slower diffusion of particles through the diffusion layer of the mercury drop (Gonqalves et al., 1985). The concentration of metal adsorbed on the vessel walls a n d / o r precipitated was measured by the method described in Part I of this work (Botelho et al., 1994). pH values were determined with an ORION ROSS electrode before metal ion was added and at the end of the measurements. The pH re-

C.M.S. Botelho et al. / The Science of the Total Enviornment 151 (1994) 101-112

mained constant ( + 0.05 pH) during the titration procedure• Zn(II) desorption. During titrations of Pb(II) solutions with particles a simultaneous Zn(II) desorption was observed. To test the influence of Pb(II) adsorption on Zn(II) desorption, solutions containing only KNO 3 and buffer were titrated with increments of particle suspension. Labile and free metal concentrations were followed by DPASV at the same voltammetric conditions previously described for titration of the Pb(II) solution. The concentration of metal ions previously adsorbed on the natural particle surfaces was determined after acidifying suspensions to pH---4, &

400

& 50.0

10.0 "8,0

II*

!! 3 0 0 ' "6.0

u

200

100

these values have been included in values of the binding capacities and mass balances. Measurements were made after 20-30 min stirring and current intensities from free and labile metal (Pb, Zn) species were measured. Simultaneous Pb(II) and Zn(II) adsorption. The adsorption of Pb(II) in the presence of a high concentration of dissolved Zn(II) was studied by titrating a 4.4 × 10 - 7 M Pb(II) and 4.4 × 10 - 7 M Zn(II) solution with a suspension of particles B (1-40 mg/1 in the final suspension) at constant pH (pH = 8.2) in a 0.01 M KNO3 and 9 × 10 -4 M acid boric/8 × 10 -5 M sodium borate medium. Organics desorption from surfaces. For particles B, a high desorption of organics from the surfaces

Corg(mg/g) P(mg/g)

E]

A

"2,0

10.0'

0.0

0.0

Foz

0""-

Cu(mgig)

N

Pb(mg/g)

•~

Zn(mg/g)

20,0

0.50

/j

30.0 20,0"

Fe(mg/g) Mn(mg:g)

[]

40.0"

"4,0

!

BragaLouro

105

i

i

• 0.40

N M

•0.30

J m

0.20

0. I0

i

Braga Louro

Foz

N • ~

logKdPb logKdCu logKdZn

7,00"

15.0

6,00"

I0.0

5,00

5,0 '

4,00

/ 0.0

w

~

Braga Louro

v

Foz

3,00 Braga

Louro

Foz

Fig. 1. (a, b, c and d). Composition of the suspended particles from the River Este at the three sampling sites of Braga, Louro and Foz (collected from 0 . 4 5 / z m m e m b r a n e filtration).

106

CM.S. Botelho et al. / The Science of the Total En~iornment 151 (1994) 101-112

was observed by the shift on the peak potential during titrations described in 3.4. To test this desorption, different concentrations of suspensions of particles B were prepared in a 0.01 M KNO 3 medium and pH was kept constant (pH--8.2) with the acid boric/sodium borate buffer. Solutions were sometimes stirred to approximate voltammetric experimental conditions. After ~ 12 h, suspensions were filtrated (0.45 /~m) and titrated with a 5 × 10 -5 M Pb(II) standard solution. Free and labile metal concentrations were followed by DPASV under the same conditions as above.

with C_ = metal concentration in particles (mol/kg~ and C w = metal concentration in soluble phase (mol/l). The obtained K d values are plotted in Fig. ld. In general, K d values are in the order Pb = Cu > Zn, in accordance with stability of hydroxide/hydroxyles and carbonates/carboxilates complexes with those metal ions. Total concentrations of lead and copper correlate with the Fe content in particles. This result confirms the importance of iron oxides in scavenging metals from solution (Stumm and Morgan, 1970).

4. Results and discussion

4.1. Characterization of the particles Clear differences in the particle composition were found at the three sampling sites as can be seen in Fig. 1, which reflect the influence of sewage and wastewaters discharge in Braga (B) and the self depuration of the river water in Louro (L) and Foz (F). Comparing the concentration of organic carbon with phosphorus in the three sites, it is observed that particles F have the highest concentrations and a ratio C / P that is within the range of values obtained by Miiller and Sigg (1990) for particles from an eutrophic lake in Switzerland. The values are characteristic of biological material (Sigg, 1987). Higher ratios for C / P in particles B and L were found, suggesting a non-biological origin for the organic carbon. Results suggest that the situation tends to natural conditions along the river although there are high artificial conditions in Braga. Total metal concentrations of lead, zinc and copper in particles increase along the river contrary to the evolution of total dissolved metal concentrations (Part I of this work). This suggests that dissolved complexes in Braga dominate the speciation of metals despite the high concentration of suspended particles. A better illustration of the distribution between water and particulate phase is given by distribution coefficients, calculated as: CO Kd = Cw

4.2. Voltammetric experiments Laboratory experiments with filtered particulate matter are used in order to quantify the binding of lead and zinc to the particle surfaces. Titration curves ip v s . Cpart are plotted in Fig. 2a. The lead in solution decreases with the addition of particles, as a consequence of the continuous ion removal by the increasing number of active sites in solution. For particles B the evolution of Pb(II) peak potential shows an abrupt shift after the third addition of particles, being constant at the end of titration. This result suggests the existence of Pb(II) labile complexes that, in principle, can be due to the metal ion adsorbed on particles or to the organic desorption from surfaces. In this case, a desorption isotherm which reaches a flat region can be attained and can explain the constancy of the peak potential for highest concentration of particles. To study the behaviour of possible desorbed organics for Pb(II) complexation a voltammetric titration without particles was performed in accordance with the procedure described in the experimental part. Straight lines with lower slopes than calibration plots have been obtained from titrations with Pb(II), which is due to the lower diffusion coefficients of dissolved complexes than those of free metal or hydroxide species. Peak potential evolution during titration of those solutions has a behaviour that is similar to the one obtained in the presence of particles, suggesting that the same kind of complexes are involved. Complexometric capacities and mean stability

C.M.S. Botelho et al. / The Science of the Total En~ornment 151 (1994) 101-112

40.0"

[71

I

30,0'

a.

0%

20.0

o •D o o

I0.0

• [30 o

0.0 0.0

•

10.0

o

o

0

i

40.0

20.0 30.0 Cpart(mg/I)

36° N "3801_

-400 qL~A°

arL

171

. m

m

-420'

eL

-440"

t~

l.d

D []

60

•

0

-460"

0

0

0

0

0

0

0

10.00

20100 30.00 Cpart(mll/I)

affinity of suspended matter for metals complexation (Stumm and Morgan, 1970). From the complexometric capacities of particles for Pb(lI) and complexing concentrations of desorbed organics, the equilibrium constant for organics adsorption on particle surfaces of sample B has been calculated. A value of logKs = 6.0 was obtained which can be compared with values presented by Buffle (1988) for fulvic acids adsorption on iron oxide surfaces. Despite the low iron content of particles from Braga, this concentration is probably enough to assume that organics are desorbed from the iron cores. Therefore, once it was proved that the shift of peak potential was due to organics in solution, desorbed from particles, results from the titration of metal solutions with suspensions of particles were corrected for complexation in solution, and interpreted in terms of binding capacities and conditional stability constants according to theory.

0

-480

.00

107

40.00

Fig. 2. Evolution of (a) peak current and (b) peak potential for the titration of a Pb(II) standard solution with suspensions of particles. Braga (O), pH = 8.2; Louro (A), pH = 7.7; Foz (O), p H = 7 . 4 . 1, Pb(II) peak potential in a 0.01 M KNO 3 solution at p H = 3.7; 2, Pb(II) peak potential in a 0.01 M KNO 3 solution at p H = 8.2 with acid b o r i c / b o r a t e buffer; 3, Pb(II) peak potential in a 0.01 M KNO 3 and 0.001 M NaHCO 3 solution at pH = 7.7 and pH = 7.4.

constants have been determined from these resuits according to expressions presented in Part I. The values obtained for the average stability constants are 9.8 < logK, < 10.2 in accordance with those obtained for Pb(II) surface complexation (Table 1) and Pb(II) soluble complexes in the river water (Table 1 in Part I), suggesting that the same kind of-complexants exists both in solution and at particle surfaces. The difference between labile and inert complexes results mainly from the difference between diffusion coefficients of the surface and dissolved species due to the size of the species and/or some stereochemical factors involved. In natural systems, organics in solution adsorb on inorganic or biogenic surfaces increasing the

The treatment of titration data by Scatchard and Ruzic methods (Eq. 7-9) shows that we can interpret results, for sample B with the existence of one predominant active site to Pb(II) adsorption. For samples L and F results suggest that two different active sites are at least responsible for Pb(II) adsorption. In Fig. 3 Pb(II) adsorption isotherm for sample F is plotted as an example and the existence of two different active sites is clear from the shape of the curve, where two flat regions are shown to exist. Table 1 Average stability constants (Kvb e KZn ) for surface complexes, complexometric capacities (Cc) and fraction of active sites with predominant affinity for the metal ion (IF e ZF) for sampling sites at Braga, Louro and Foz

IoglKpb (l/mol) Iog2Kpb (l/mol) CCpb (mol/kg) 2l F eb Feb log 1Kzn (l/mol) log2 K.zn (l/mol) Cczn (mol/kg) I Fz n 2Fz n

Braga

Louro

Foz

10.3 -0.077 --6.79 5.79 0.017 0.59 0.41

10.5 9.0 11.023 0.26 0.74 8.5 -0.021 ---

10.3 8.9 0.038 0.28 0.72 8.3 -0.026

108

CM.S. Botelho et al. / The Science of the Total Entiornment 151 (1994) 101-112

metal as those obtained in solution. For these samples, if we consider that the same kind of complexants exist in the river solution and on particle surfaces, results are not very much affected by this simplification. Evolution of peak potential during titration with particles F shows a regular shift only at the end of titration, which probably indicates that organic desorption begins to be significant for this sample only for high particle concentre.tions. It is worthwhile to emphasize that the change of IogD/DM can compensate the variation of logK[L] in the DeFordHume equation and AE -~ 0. Due to the complexity of the systems, only for sample L has the differential parameter been calculated and the differential equilibrium function been defined. The average heterogeneity of surface complexants, calculated from the slope of the differential equilibrium function (F = 0.38), is in accordance with the value obtained for river

0.040"

0.030 O

0,020

¢~

0.010

0,000 0.00e+0

•

,

•

2.00e-9

i

.

,

4.00e-9 IPbItM)

6.00e-9

8.00e-9

Fig. 3. Adsorption isotherm for Pb(II) on particle surfaces from the sampling site Foz (F). Experimental points and calculated isotherm.

Titration results for samples L and F have been corrected for complexation in solution using the DeFord-Hume method and assuming identical diffusion coefficients for free and complexed 60.0"

30,0"

[]

50.0"

+

4- 4000

'~ 4 0 . 0 " 0 ~ 30.0"

+

O

0

O

4-

320.0"

0

g~

gl.

4-

- 10.o.

+r~ O

4-

tp

10.0" 0.0

0 . 0 ~ ,~ 0.0

!

0.0 0.02

[]

4-

0

-- 2 0 , 0 "

4n

20.0 Cpart(mg/I)

40.f

!

40.0

20.0 Cpart (mlt/I)

"

0.03

rn

N O

o ~

+

o+ "U

t~

a

4-'~

~0.01

44-

~0. 0 1 t~

O

0.00 ~ O,Oe+0

b

0

•

!

2.5e-8 lZnl(M)

5.0e-8

0.00 O.Oe+O

2.5e-8 [Znl(M)

5.0e-8

Fig. 4. (a) Titration curves and (b) adsorption isotherms for Zn(II) adsorption on particle surfaces from Braga (O) and Foz (rT). Titration without Pb(II) in solution (+).

109

C.M.S. Botelho et al. / The Science of the Total Enviornment 151 (1994) 101-112

solution at the Louro site (F = 0.40) (see Part I of this work). It must be pointed out that the average stability constants calculated from the isotherm are within the variation range of the differential parameter (7.5 < IOgKDEF < 10.5). The concentration of metal ions already adsorbed on the natural particle surfaces has been measured by desorption at pH 3.7. Results indicated that ~ 70% of total Pb in particles B has been desorbed at pH 3.7, whereas only 40% for samples L and F are at the surface. This suggests that particles B, most of them coming from domestic sewage, adsorb a large quantity of Pb ions on contact with industrial wastewaters. Samples L and F have identical properties, different to B, indicating a similar nature of those particles, their order of magnitude being the same as those obtained by Miiller and Sigg (1990) for natural particles. From adsorption results, it is possible to conclude that particles B have the highest complexometric capacity which can be due to a higher surface area. These particles, behave in suspension as flakes with high porosity, which can also account for the high organics desorption during titration experiments. Particles F have more capacity to complex Pb ions than particles L, probably because of the higher pH value and the biological nature of organic carbon. Average stability constants obtained for samples B and F agree with the values determined for organics in solution (Botelho et al., 1994). A similar result was reported by Gonqalves et al. (1987) and (1989) for Cu(II) on the Klebsiella p n e u m o n i a bacteria and Selenastrum Capricornuturn Printz alga surface and dissolved exudates. During experiments with Pb(II) desorption of Zn(II) ions was observed. In order to verify if this was produced by the simultaneous Pb(II) adsorption, a blank, at the same pH, was titrated with suspended particles. Results are plotted in Fig. 4 for B and F samples as an example (sample L behaves like sample F) and the calculated parameters presented in Tables 1 and 2. Only particles B present an additional Zn desorption in the presence of Pb ions. Samples L and F have the same behaviour, an identical Zn desorption with or without Pb in solution.

Evolution of peak potential values are the same for both desorption cases. As happens for titration of Pb solution with particles B, a high shift on Zn peak potential was observed for the first points. In this set of titrations metal ion is added to a suspension of particle surfaces and so, after the cathodic shift there is an anodic one typical of a titration of constant complexing concentration with metal ions, where labile complexes are being formed. For sample B, in the presence of Pb(II), zinc adsorption data are well fitted to a Scatchard or a R~zic plot with two predominant active sites. In the absence of lead, the zinc adsorption isotherm is of the Langmuir type showing the existence of only one kind of active site for all samples. The stability constant for zinc in sample B is higher than the two values obtained for zinc surface complexes in the presence of Pb indicating a competition between Pb and Zn (Tables 1 and 2). Stability constants for Zn surface complexes are similar for samples L and F, and there is a good agreement between values obtained from the two sets of desorption experiments. The results show that conditional stability constants of Pb and Zn with natural particles are similar for the three sampling sites (except for Zn on sample B, in the presence of Pb), but there is a clear difference between conditional constants for lead and zinc, with higher stability constants for lead in all cases. This is in agreement with results on the binding of lead and zinc to well defined particle surfaces (Shindler and Stumm, 1987) and with the tendency of these metal ions to bind to oxygen ligands. The same result was obtained by Mi~ller and Sigg (1990) for lead and zinc complexation on natural particle surfaces. Conditional stability constants for adsorption Table 2 Average stabilityconstants (Kzn) for surface complexesand complexometriccapacities (Cc) obtained from Zn(II) desorption data withoutPb(II) in solution

logKzn(l/mol) Cczn (mol/kg)

Braga

Louro

Foz

8.0 0.017

8.3 0.021

8.1 0.027

110

CM.S. Botelho et al. / The Science of the Total Enciornment 151 (1994) 101-112

of lead and zinc on particle surfaces from Foz are similar to those obtained by Miiller and Sigg (1990) who studied metal complexation on particle surfaces from a less polluted river. Complexometric capacities obtained in this work (sample F) are approximately three times as high as the values presented by these authors. The difference can arise from the higher iron and organic carbon content of sample F. For sample B the conditional constants obtained for Pb are outside the range of values presented in the above paper and the binding capacity is 10 times as high. This result shows the different nature of these particles. Since they have a similar content in organic carbon and a lower concentration of iron, the higher binding capacity of particles from Braga can only be explained by a higher surface area. For zinc the binding capacity is also higher, but the value of the stability constant obtained from the experiment without lead is within the range of values presented by the above authors. It seems that, despite the different nature of particles B, the same type of groups with similar affinities are involved in the zinc complexation on these polluted particles and on less polluted surfaces. For samples L and F, stability constants for complexation of lead with surface groups are also higher than values obtained by Miiller and Sigg (1990), which suggests some influence of the anthropogenic organics from Braga. For zinc, the perfect agreement found for results from the two different kinds of particles may indicate the presence of the same type of groups for this cation. A titration of a Zn and Pb solution with Cvb = Czn = 4.4 × 10 -7 M, was also made with a suspension of particles B to understand what happens with Pb adsorption in the presence of the same concentration of soluble zinc, as observed in the river. Complexometric parameters were obtained considering a competition for surface sites between lead and zinc. From the data a total complexometric surface capacity of 0.094 + 0.001 mol/kg and average stability constants with the values logK.pb = 10.4 and logKzn = 6.9 have been obtained. The results are in perfect agreement with those

obtained without Zn, the difference being in the fraction of sites occupied by each metal ion. Lead occupies 80% of total active sites when there is no Zn in solution, but in the presence of this metal ion at the same concentration it occupies only 50% of the total surface active sites, suggesting a competition of both cations for this type of sample. The combination of field measurements and laboratory experiments allows a check for the distribution of metal ions between particulate and dissolved phase which can be explained in terms of simple interactions with surfaces and speciation in solution, on the basis of the complexometric parameters obtained. For this purpose the distribution coefficients are derived from speciation calculations which include the interactions with suspended particles and dissolved organics by using the binding parameters given in Table 3 of Part I and in Table 1 of this work, Zn concentration being roughly 1000 times higher than Pb concentration. Zn and Pb competition for dissolved complexants has also been considered despite the fact that Zn complexometric parameters in solution have not been determined. However, we have assumed that total concentrations of dissolved organic complexants are those found for lead complexation and that stability constants for soluble Zn complexes are the same for surface complexation (Figs. 5 and 6). It was observed that zinc speciation in solution does not influence the Kd value for zinc. Whether Pb and Zn competition is or is not considered for dissolved ligands, speciation of the last cation is always controlled by free metal concentration. Calculated K s values for Zn agree, in a general way, with field values as can be noticed in Fig. 5. An exception was found for sample L, with a higher value for Kd obtained from field data. This can be due to errors of the metal adsorbed concentration in the particles used in K d values determined in the field, since particulate metal concentration was obtained from total metal concentration measured by AAS on digested sampies. On the other hand calculated K d values were obtained using surface complexometric

C M.S. Botelho et al. / The Science of the Total Enviornment 151 (1994) 101-112

6.0e-8

7.0 6.0

111

a Pb(H)I Zn(ll) ]

l

J

[•

f

L

i orlinic Pb2+ s particles

4.0e-8

clrbonates hydroxides

2.0e-8

O,Oe+O 3,0 3.0

Braga

4.0

5,0

6.0

Louro

Foz

7.0

Kdexp

Fig. 5. Comparisonof distribution coefficientsmeasured in the River Este with calculated values using experimentally obtained binding constants, adsorption capacities, concentrations, pH, etc., correspondingto the conditionsdeterminedin the river. parameters, and therefore taking into account only metal surface concentrations. Measurement of Zn desorbed at pH = 3.7 from particles L shows that only 8% of the total metal content is on the surface. Thus, if K d from field measurements is only the surface metal concentration, the agreement with the calculated value is perfect. In terms of lead, adsorption capacities for lead are always higher (5-20 times) than total metal concentration. Because of the higher Zn concentration in the river, a competition of the two metal ions for surface sites is playing a role. So K d values for lead are very sensitive to solution composition and especially to soluble zinc speciation. Although speciation calculations have the limitations pointed out above, from Fig. 6 it can be concluded that free lead is only detectable in L and F samples and that total lead concentration has the highest value in B due to the wastewater inputs. This station has a higher percentage of soluble lead bound to organics than adsorbed on particles, which is inverted in L sample due to water collection for domestic supply, it being reequilibrated at F sampling point. Due to the higher concentration of zinc and its smaller affinity for most organic ligands, the free cation dominates speciation in all stations and inorganic complexes are more important than for

"4.0e-6 6.0e-5

orsauies particles carbonates hydroxides

4,0¢-5

"2.0e-6

i

2.0e-5 O.Oe+O Braga

Louro

Foz

).Oe+O

Fig. 6. (a) Pb(II) and (b) Zn(II) speciation calculated with experimentallyobtained complexometricparameters. lead. Organic speciation of soluble species is more important in L and F samples. 5. Conclusions

Adsorption experiments with natural particle surfaces confirm, as shown by Gon~alves et al., 1985, 1987; Miiller and Sigg, 1990, 1991; Xue et al., 1985, that a surface complexation approach can be used for metal adsorption on surfaces. Differential anodic stripping voltammetry has been shown to be able to distinguish between dissolved metal ions and particulate bound metal ions for metal and particle concentrations close to those in the natural environment. Conditional adsorption parameters can be derived from such experiments by a simplified approach and give consistent results for suspended matter and filtrate solutions (Part I) collected at the same sampling site. For the highly polluted sampling site, the high anthropogenic dissolved organic content largely

112

CM.S. Botelho et al. / The Science of the Total Enviomment 151 (1994) 101-112

controls lead speciation. Nevertheless, results for this site suggest that dissolved complexants are macromolecules and polymers, perhaps also colloids, that aggregate and sediment between Braga and Foz, scanvenging metal ions and anthropogenic organic matter from solution to sediments. Comparison of the distribution coefficient of Zn and Pb calculated under field conditions with the experimental results indicates that the distribution of these metals is in agreement with the experimental complexometric parameters for the surface and soluble species. Competition of the different metal ions for the available surface sites and dissolved complexants has been noticed. Distribution of metal ions between suspended particles and solution can be predicted by a speciation model treating surface ligands as dissolved complexants. Mean values of stability constants are found to be a good approach for complex affinity, so the estimated values seem to be close to the real ones. From our results it is possible to predict changes in dissolved and particulate concentrations of lead and zinc due to the environmental factors. References APHA, AWWA, WPCF (Eds), 1985. Standard Methods for the Examination of Water and Wastewater. 16th edn. Berg, C.M.G. van den and J.R. Kramer, 1979. Determination of complexing capacities of ligands in natural waters and conditional stability constants of the copper complexes by means of manganese oxide. Anal. Chim. Acta, 106:113-120. Botelho, C.M.S., R.A.R. Boaventura, M.L.S.S. Gon~alves, 1994. Interaction of lead(II) with natural river water: part I. Soluble organics. Sci. Total Environ., in press. Buffle, J., 1988. Complexation Reactions in Aquatic systems - - an Analytical Approach. Ellis Horwood, U.K.

De Ford, D.D. and D.N. Hume, 1951. The determination of consecutive formation constants of complex ions from polarographic data. J. Am. Chem. Soc., 42: 5321-5322. F6rstner, U. and G.T.W. Wittmann, 1981. Metal Pollution in the Aquatic Environment. Springer-Verlag, 2nd edn. Gon~alves, M.L., L. Sigg and W. Stumm, 1985. Voltammetric methods for distinguishing between dissolved and particulate metal ion concentration in the presence of hydrous oxides. A case study on lead(II). Environ. Sci. Technol., 19: 141-146. Gon~alves, M.L.S., L. Sigg, M. Reutlinger and W. Stumm, 1987. Metal Ion Binding by Biological Surfaces: Voltammetric Assessment in the Presence of Bacteria. Sci. Total Environ., 60: 105-119. Gon~alves, M.L.S. and A.C.L. Concei~o, 1989. Metal ion binding of copper(II), zinc(II) and lead(II) by alga Selenastrum Capricornutum Printz, Sci. Total Environ., 78: 155-166. Lee, J., 1983. Complexation analysis of fresh waters by equilibrium diafiltration. Water Res., 17: 501-510. Miiller, B. and L. Sigg, 1990. Interaction of trace metals with natural particle surfaces: comparison between adsorption experiments and field measurements. Aquatic Sciences, 52/1: 75. Rfizic, I., 1982. Theoretical aspects of the direct titration of natural waters and its information yield for trace metal speciation. Anal. Chim. Acta, 140: 99-113. Shindler, P.W. and W. Stumm, 1987. The surface chemistry of oxides, hydroxides and oxide minerals. In: W. Stumm (Ed.), Aquatic Surface Chemistry. Wiley-Interscience, New York, pp. 83-110. Sigg, L., 1987. Surface chemical aspects of the distribution and fate of metal ions in lakes. In: W. Stumm (Ed.), Aquatic Surface Chemistry. Wiley-lnterscience, New York, pp. 319-349. Stumm, W. and J.J. Morgan, 1970. Aquatic Chemistry - - an Introduction Emphasizing Chemical Equilibra in Natural Waters. Wiley-Interscience, New York. Tessier, A., R. Carignan, B. Dubreuil and F. Rapin, 1989. Partitioning of zinc between the water column and the oxic sediments in lakes. Geochim. Cosmochim. Acta, 53: 1511-1522. Xue, H.B., W. Stumm, L. Sigg, 1988. The binding of heavy metals to algal surfaces. Water Res., 22: 917-926.