Effects of colloids on metal transport in a river receiving acid mine drainage, upper Arkansas River, Colorado, U.S.A.

Applied

Geochemisrry,

Vol. 10, pp. 285-306, 1995 Elsevier Science Ltd Printed in Great Britain

088%2927(95)00011-9

Effects of colloids on metal transport in a river receiving acid mine drainage, upper Arkansas River, Colorado, U.S.A.

Briant

A. Kimball

U.S. Geological Survey, 1745 W 1700 S Rm 1016, Salt Lake City, UT 84104, U.S.A.

Edward

Callender

U.S. Geological Survey, MS 432 National Center, Reston, VA 22092, U.S.A.

and Ellen

V. Axtmann

U.S. Geological Survey, 3215 Marine St, Boulder, CO 80303, U.S.A. (Received 24 February 1994; accepted in revised form 1 March 1995)

Abstract-Inflows of metal-rich, acidic water that drain from mine dumps and tailings piles in the Leadville, Colorado, area enter the non-acidic water in the upper Arkansas River. Hydrous iron oxides precipitate as colloids and move downstream in suspension, particularly downstream from California Gulch, which has been the major source of metal loads. The colloids influence the concentrations of metals dissolved in the water and the concentrations in bed sediments. To determine the role of colloids, samples of water, colloids, and fine-grained bed sediment were obtained at stream-gaging sites on the upper Arkansas River and at the mouths of major tributaries over a 250-km reach. Dissolved and colloidal metal concentrations in the water column were operationally defined using tangential-flow filtration through 0.001~pm membranes to separate the water and the colloids. Surface-extractable and total bed sediment metal concentrations were obtained on the <60+m fraction of the bed sediment. The highest concentrations of metals in water, colloids, and bed sediments occurred just downstream from California Gulch. Iron dominated the colloid composition, but substantial concentrations of As, Cd, Cu, Mn, Pb, and Zn also occurred in the colloidal solids. The colloidal load decreased by one half in the first 50 km downstream from the mining inflows due to sedimentation of aggregated colloids to the streambed. Nevertheless, a substantial load of colloids was transported through the entire study reach to Pueblo Reservoir. Dissolved metals were dominated by Mn and Zn, and their concentrations remained relatively high throughout the 250-km reach. The composition of extractable and total metals in bed sediment for several kilometers downstream from California Gulch is similar to the composition of the colloids that settle to the bed. Substantial concentrations of Mn and Zn were extractable, which is consistent with sediment-water chemical reaction. Concentrations of Cd, Pb, and Zn in bed sediment clearly result from the influence of mining near Leadville. Concentrations of Fe and Cu in bed sediments are nearly equal to concentrations in colloids for about 10 km downstream from California Gulch. Farther downstream, concentrations of Fe and Cu in tributary sediments mask the signal of mining inflows. These results indicate that colloids indeed influence the occurrence and transport of metals in rivers affected by mining.

INTRODUCTION Acid mine drainage contaminates many streams in the mining districts of the Rocky Mountains (Colorado Department of Health, 1988; Moore and Luoma,

1990). Drainage from inactive and abandoned mines often enters upland watersheds, and eventually mixes with non-acidic water. Mixing with downstream tributaries causes dilution and neutralization of the acid and the formation of Fe colloids (Runnells and Rampe, 1989; Kimball and Wetherbee, 1989). Colloids are solids with effective diameters in the size range from about lo-’ to 10m6 m. They are also defined as solids that do not settle from the aqueous phase (van Olphen, 1977). The small size results in extensive surface area that strongly influences the

partitioning of toxic metals through sorption and coprecipitation (Jenne, 1977; Stumm and Morgan, 1981; Morel and Gschwend, 1987). Most studies of partitioning between dissolved and particulate phases have relied on filtration through 0.45;um membranes. This is adequate for solutes that tend to partition to the aqueous phase and in systems without Fe colloids. Jones et al. (1974) showed that metals in surface water, like Al, Fe, and Mn, require filtration with pore-size membranes smaller than 0.45 ,um to define dissolved concentrations. In Ferich systems, even 0.1~pm membranes may allow Fecolloids to pass through and be analyzed as part of the dissolved phase (Kimball ef al., 1992). Ultrafiltration separates much smaller particles from the aqueous phase (Josephson, 1984; Hemandez and Stallard, 285

286

B. A. Kimball et al.

1988). This paper presents the results of a study using tangential-flow ultrafiltration for samples in a river affected by acid-mine drainage. By defining the colloidal fraction with ultrafiltration, it is possible to assess the role of colloids in metal transport. Much work has been done to understand the processes affecting metals in natural systems receiving acidic mine drainage (Barnes and Clarke, 1964; Bencala and McKnight, 1987; Bigham et al., 1990; Chapman et al., 1983; Davis, 1988; Filipek et al., 1987; Honeyman and Santschi. 1988; Kimball et al., 1990; McKnight et al., 1988; McKnight and Bencala, 1989; Moore et al., 1991; Nordstrom. 1985). Most studies, however, have not specifically isolated colloidal fractions to investigate the role of colloids in metal transport (deGroot et al., 1982; Perhac and Wheelan, 1972; Forstner and Muller, 1973; Andrews, 1987; Marron, 1989; McKallip et al., 1989; Axtmann and Luoma, 1991). Hoffman et al. (1981) distinguished between dissolved and colloidal concentrations of metals in isolated organometallic complexes from the Mississippi River and Yan et al. (1992) evaluated metals in colloids in estuarine waters. Others have looked at the partitioning in sediments and colloids of the Mississippi River (Taylor et al., 1990; T. F. Rees and J. F. Ranville, U.S. Geological Survey, pers. commun., 1990; Marley et al.. 1991). Davis et al. (1991) describe partitioning of metals between 0.45;um filtered water and suspended particulate phases in Clear Creek, Colorado, which receives acidic mine drainage. Material passing through a 0.45;um membrane, however, may not represent truly dissolved phases in streams affected by acid mine drainage (Kimball and McKnight, 1989). In 1983 and 1985, surges of mine discharge from the Yak Tunnel in California Gulch released several thousand liters of metal-rich water to the Arkansas River. These surges produced distinct clouds of ochre-colored colloids which were visible for about 100 km downstream from Leadville. Cores of bottom-sediment from Pueblo Reservoir obtained in October 1987 showed distinct ochre layers, indicating rapid transport and sedimentation of these colloids (Callender et al., 1989). Even under normal conditions colloids are visible in the discharge from California Gulch, but the colloids disperse within a few hundred meters. This study focuses on the occurrence and transport of colloids in the Arkansas River and their influence on the transport of metals downstream to Pueblo Reservoir. Objectives for sampling water and sediment from the Arkansas River were (1) to find the extent of colloid transport downstream from the Leadville area, (2) to determine the chemistry of colloids, and (3) to study the effect of colloids on the transport of metals. Iron, Mn, and Al were included in the study because they commonly form reactive surface coatings (Jenne, 1977; Balistrieri and Murray, 1982; Benjamin, 1983; Schultz et al., 1987; Honeyman et al., 1988). Cadmium, Cu, Pb, and Zn were included because of their toxic effects on aqua-

tic organisms (Forstner Gough ef al., 1979).

and

BASIN AND WATER-QUALITY

Whittmann,

1979;

CHARACTERISTICS

Metal-rich water enters the headwaters of the Arkansas River from mines and tailings piles in the Leadville, Colorado, area (Fig. 1). Part of this area is classified as a national priority site for cleanup of metal contamination. The upper Arkansas River basin begins at the continental divide and includes the highest peak in Colorado, Mt. Elbert at 4399 m. We studied the Arkansas River in a 250 km reach from upstream of Leadville to Pueblo Reservoir (Fig. 1). Samples were taken at 14 river sites and 13 major tributaries during low flow conditions of October 1988, and at 10 river and 6 tributary sites during snowmelt runoff in May 1989. Instantaneous discharge was measured for each site and water samples were collected by a discharge-integrating method (Ward and Harr, 1990). The river above Pueblo, Colorado, which is at 1423 m. drains about 5940 km’ (Crouch et al., 1984). Mean annual discharge at Canon City (river km 197) was 19.8 m3 s-’ for water years 1962-1979. Variation in streamflow is strongly seasonal; during spring and early summer, snowmelt runoff dominates the discharge of the Arkansas River [Fig 2(a)]. Discharge can increase tenfold from Leadville to Pueblo Reservoir [Fig. 2(b); Table 11. Values of discharge during the two sampling periods are low compared to October and May stream flow during the period 1981-1989 (U.S. Geological Survey, 1981-1989). Between river kilometers 20 and 25 and also between 183 and 194 km there are diversions of water from the river, mostly for irrigation (Abbott 1985). The diversions between river kilometers 183 and 194 are large enough to affect the calculations of instream metal loads, and so the loads will only be considered through the site at 183 km. During snowmelt runoff in May 1989, pH of the Arkansas River and its tributaries generally was lower than during low flow in October 1988 (Table 1). Tributary pH values strongly affect the river pH during runoff because of the lower pH and lower buffering capacity of snowmelt. Downstream from the headwater tributaries and during low-flow conditions, pH values were higher. Major solutes in water from the Arkansas River reflect the weathering of many kinds of rocks in the basin. Inflows affected by acid mine drainage in the Leadville area reflect the oxidation of metal-sulfide deposits, producing acidic, sulfate-rich water. Some of this water, however, is neutralized by flow through carbonate rocks before discharging from the Leadville Drainage Tunnel. Discharge from the Yak tunnel and tailings piles in California Gulch is neutralized by mixing with alkaline discharge from the Leadville Sewage Treatment Plant before entering the Arkansas River (Table 1). Once neutralized, the

Metal transport in an acid mine impacted

river

287

ST KEVIN GULCH

EXPLANATION

L-_--_

1’4

Sampling&e-number

A32

Gagingrlation-number

represents kilometers downstream from Climax Colorado represents kilometers downstream from Climax Colorado

__--.

1

0

lil

, 10

2;

3/,KILOMETER;

20

30 MILES

Fig. 1. Map showing the upper Arkansas River basin above Pueblo, Colorado. Sampling locations are indicated by symbols and the downstream distance of the site in kilometers from Climax, Colorado, which is upstream from Leadville.

resulting water generally is a calcium, magnesium, sulfate, bicarbonate water. Inflows from the western side of the Arkansas River, downstream from the Leadville area, generally drain areas of igneous and metamorphic rock. These contribute calcium, sodium, and bicarbonate to the river. Tributaries on the eastern side of the Arkansas River also drain areas of igneous and metamorphic rock, but also areas of sedimentary rock (Crouch et al., 1984). The chemistry of Badger Creek (129 km), downstream from Salida, reflects the weathering of a shale containing halite (Table 1). The chemistry of inflows from Tallahassee (182 km), Fourmile (202 km), and

Beaver (223 km) Creeks is strongly influenced by weathering of a shale that contributes calcium, sodium, and sulfate. Silica also is reported in Table 1 to indicate the overall silicate weathering in the basin.

METHODS Ideally, the concentration of a metal in a distinct geochemical phase can be determined. The rigors of sampling natural materials, however, require that phases be operationally defined. A geochemical phase is a physically and chemically homogeneous substance like a solid mineral

B. A. Kimball et al.

288

_ (a) MEAN

DAILY DISCHARGE

100

,

,

,

,

,,

7/l I09

4/1/09

111169

1O/1/88

I,

I,

DATE I,

I,,

I

I

I

I,

13

I (b) INSTANTANEOUS DISCHARGE

-Q-

LOW FLOW, OCTOBER 1988

--m-HIGH

0.1

FLOW, MAY 1989

“‘~““““““““’ 0

100

50

150

200

DISTANCE, IN KM

Fig. 2. (a) Mean daily discharge of the Arkansas River, near Leadville, Colorado for the 1989 water year (b) Instantaneous discharge at sampling sites during October 1988 and May 1989.

phase, a surface-coating phase, a suspended colloidal phase, or an aqueous phase (Stumm and Morgan, 1981). Traditionally, the “dissolved” fraction in an aqueous phase is defined by passing the water sample through a 0.45pm filter. The suspended particulate fraction is defined as everything greater than 0.45 pm. For Fe-rich streams affected by mining, 0.45 pm is neither an effective nor a natural break point for measurement of dissolved and particulate concentrations. The chemical behavior of Fe, even at pH near 3.0 favors the formation of Fe-rich colloidal material that can be much smaller than the 0.45ym cutoff. Our working definition of the “dissolved” fraction relies on tangential-flow ultrafiltration through a membrane with an effective pore size less than 0.001 ,um (Hernandez and Stallard. 1988).* For comparison to the ultrafiltrate, water samples also were filtered through O.l- and 0.45~pm membranes using tangential-flow filtration with a Millipore Minitan.? The same filter apparatus was used with identical pump settings for all samples, only changing membrane *Commercially these filters are sold as 100,000 molecular-weight nominal pore size, or 100,000 daltons pore size. tThe use of trade and brand names does not imply endorsement by the U.S. Geological Survey but is for identification purposes.

sizes, to minimize artifacts of filtration (Horowitz ef ul., 1992). Moran and Moore (1989) used a lO.OOO-molecularweight nominal pore-size membrane with a cross-flow process to evaluate “truly dissolved” aluminum and organic carbon compounds in seawater near Nova Scotia. Because of aggregation, Fe colloids are present in a continuum of sizes ranging from about 0.001 pm to greater than 0.45 pm. Suspended sediment in the upper 100 km of the Arkansas River mostly consisted of aggregated colloids rather than the normal sand, silt, or clay of other rivers. Examination of aggregated colloids from the Leadville area, using a scanning electron microscope, showed that the O.OOl-pm poresize membranes generally were adequate to remove aggregates of Fe colloids (Ranville er al., 1989). Downstream from Badger Creek (129 km) the amount of silt and clay increased. For this study, the colloidal fraction included everything greater than 0.001 pm. Suspended solids greater than 60 ,um were excluded from the colloidal fraction by filtering through using a nylon screen before tangential-flow filtration. The chemistry of the colloidal fraction was determined by direct and indirect measurements. Direct measurement of metal concentrations in the colloidal fraction was made on about 1 g of solid sample that was separated from the water by dewatering about 100 I of water using a Millipore Pelican

289

Metal transport in an acid mine impacted river Table 1. Concentrations of dissolved major

ions

in water from the Arkansas River basin Ca

Mg

Na

Cl

24 39 33 33 27 23 21 26 35 36 41 40 39 63

9.6 16 13 14 10 8.2 6.3 7.3 8.5 8.8 11 11 11 20

1.5 2 2.2 2.2 4.5 4 4 5.2 7.9 8.3 12 13 13 28

0.4 0.8 0.7 0.7 2.1 2.3 1.4 1.7 1.9 2.2 7.1 9.5 9.8 12

43 3.3 50 3.1 1.4 2.7 2.3 16 20 10 26 23 54

3.3 2.5 4 4.1 1.6 3.3 9.1 14 64 9.9 56 32 50

2.5 0.7 2.4 2.3 1.4

pH

Temp Cond

0.31 0.42 0.57 0.58 1.75 3.28 3.96 6.11 6.93 9.03 9.68 10.47 7.08 7.99

8.06 8.08 8.3 7.78 8.2 7.88 7.79 8.43 8.38 8.41 8.52 8.36 8.6 8.54

0.0 2.5 3 10.5 5.5 1.0 6.5 9 10 6.5 8 9 8.5 7

182 285 238 291 218 202 182 213 263 268 205 330 334 525

15 19 25 26 42 74 86 117 129 162 183 202 223

0.08 0.21 0.02 0.59 0.37 0.29 1.87 0.29 0.17 0.43 0.05 0.87 0.02

7.13 7.43 6.83 8.14 7.52 8.25 7.99 8.76 9.18 8.69 8.84 8.22 7.85

6 4.5 5 8 8 4 8 10.5 13.5 7 13 11 7

739 100 88 9.6 854 96 na 9.7 77 10 120 17 200 22 450 50 910 58 283 46 633 50 608 80 1602 252

Arkansas River, May 1989 East Fork, above ieadville Drain East Fork. below Leadville Drain Near Leadvillc Above California Gulch Below California Gulch Below Lake Fork Near Malta At Salida At Parkdale At Portland

14 18 20 24 25 27 32 111 183 215

2.59 2.71 8.26 6.82 6.88 6.93 10.47 43.02 42.45 36.03

7.57 7.37 7.47 7.06 7.31 7.57 7.14 6.98 8.05 8.54

9 12 7.6 8 7 11 7.5 7 15 13.7

Tributaries, May 1989 Leadville Drain Tenncssec Creek California Gulch Lake Fork Chalk Creek Fourmile Creek

15 19 25 26 86 202

0.08 4.67 0.02 1.78 4.41 0.31

6.78 6.67 6.58 7.27 7.29 7.44

7 4 16 7 7 7

Site name

Dist

Arkansas River, October 1988 East Fork, above Leadville Drain East Fork, below Leadville Drain Near Leadville Below California Gulch Near Malta At Granite At Buena Vista Near Nathrop At Salida Near Wellsvillc At Cotopaxi At Parkdale At Canon City At Portland

14 18 20 25 32 46 71 96 111 120 150 183 194 215

Tributaries, October 1988 Leadvillc Drain Tennessee Creek California Gulch Lake Fork Lake Creek Cottonwood Creek Chalk Creek South Fork, Arkansas River Badger Creek Texas Creek Tallahassee Creek Fourmile Creek Beaver Creek

Disch

SO1 HCOl SiOz

12 72 52 68 44 33 2.7 26 15 23 35 39 23 28

300 8 490 13 12 na 0?“7 17 1.5 12 84 79 7.7 2.4 26 69 11 190 11 780

109 112 103 101 79 78 71 95 135 140 155 157 159 195

na na 7.4 7.4 9.2 8.7 8.8 11 13 13 12 12 12 12

144 na 41 na 13 2.9 36 na 34 4.3 na 7.6 77 16 262 14 199 24 205 18 317 20 204 14 276 17

138 160 88 111 139

14 18 9.4 9.9 14

5.6 7 3.5 3.7 6.5

1.1 1.2 1.2 1.1 1.6

0.6 0.6 0.5 1.5 0.8

13 25 13 13 32

54 59 31 35 34

1.7 5.4 5.7 6.4 6.6

1:; 1:;

12 11 21 13

3.9 4.4 5.4 2.8

1.8 1.7 4.7 2.3

0.9 0.8 0.7 2.5

20 15 20 15

36 34 40 62

6.9 6.2 II7.9

221

28

8.6 10

3.4

47

75

13

2.2 0.4 12 0.8 0.5 17

423 5.8 485 9.4 11 490

159 13 32 18 29 198

11 5.9 14 7.1 8 14

1038 120 53 3.7 46 3.6 1.2 1 1023 98 52 15 72 5.8 1.8 2.9 76 11 1 2.6 1182 134 45 69

Dist, distance downstream from Climax, Colorado, in kilometers; Disch, instantaneous discharge, in cubic meters per second; Temp. temperature, in degrees Celsius; Cond. conductivity, in microsiemens per centimeter at 25°C: na. not analyzed; all concentrations in milligrams per liter. OM-141 tangential Row unit with the 0.001~pm pore size. The direct measurement was made at six sites, and provided reliable concentrations for each of the metals. Indirect measurement of metal concentrations in colloids was made in an unfiltered, acidified sample of water. This represented an in-bottle digestion of the Fe colloids. Metal concentrations in the dissolved fraction were subtracted from this unfiltered concentration. The indirect measurement was reliable for Fe, Mn, and Zn at each of the sites, and was used to calculate loads of the colloidal fraction for each site. Filtered and unfiltered water samples for analysis of metal concentrations were acidified in the field with doubledistilled nitric acid to a pH <2.0. Metal concentrations were then determined by inductively coupled plasma-atomic emission spectrometry (ICP-AES; Garbarino and Taylor,

1979). Copper and Pb were determined by graphite furnace atomic absorption (GFAA; Fishman and Friedman, 1989). Anions were analyzed in O.OOl-pm filtered, unacidified samples by ion chromatography (Fishman and Friedman, 1989). Reagents for analysis of ferrous iron by the 2,2’bipyridine calorimetric method were added immediately after filtration to aliquots of unacidified sample (after Brown eral., 1970). These samples were kept out of the light until ferrous iron was determined calorimetrically in the laboratory (McKnight etal., 1988). Ferric iron was assumed to be the difference between the O.OOlym filtered Fe and the measured ferrous iron concentration. Ferrous iron was detected in some samples of non-acidic streamwater of the Leadville area, and was used to calculate a redox state of Fe phases (Bail and Nordstrom, 1991). Where the concen-

B. A. Kimball et al.

290

Table 2. Concentrations of dissolved and colloidal metals in water samples from the Arkansas River basin Al

Dist Diss

Coil

Diss

IO 0.5

0.5 0.1 0.04

I 3 15 7 8 IO 13 x 7 5 2 4
1 7
3 5


30

3 2 2.4 1.4 1.2

169
1

Diss

Detection limits and orecision Equipment blank ’ Detection limit Precision Low/high Arkansas 14 18 20 25 32 46 71 96 111 120 I50 183 194 215

River, October

Tributaries, 15 19 25 26 42 74 X6 117 129 162 183 202 223 Arkansas 14 IX 20 24 25

October

1988
River, May 1989 110 <10
27

ill)

32 32 Ill 183 215


Tributaries, 15 19 25 26 86 202 Dist,

concentrations

I 0.3


Mn

Fe CON

Diss

10 0.32

4 3 1 0.07

COII

Coil

2

-

I I

-

IWO

0.3

7.2


30 8 <3 <3 3 13 3 15 4 5 6 10 4 13

120 162 158 1110 264 125 74 58 103 81 77 75 65 330

30 200 I33 813 262 115 26 14 19 I7 9 6 I 29

I.6 1.7 I.9

2 1 85.6 1
80 100 <3 100 13 24 5 3 <3 6 47 <3 8

1570 130 5790 210 13 69 262 99 42 89 <3 181 112

8 8 I 10 IO
3 3
8 6 6 7 7 9 9 3 1 3 4

1 1 3 9
26 17 37 30 30 30 30 34 IO 7 8

54 9 86 12 3
6
7 8 13 11 9 5

11
67 63 8 32 24 I5

9

1
1


2 2 3.4

I

1

1

Pb

Diss

Diss

3 0.3 0.1 0.03

1000 4.5

Zn Coil

IO 1.2

CON

Diss

5 3 1 I000 1.3 7.4

20 490 274 976 340 356 133 60 47 67 22 25 43 31

<3 <3 65 344 <3 35 14 12 15 <3 20 <3 13 <3



iI


3
20
il il 26 26 2
1400 20 20,500 30 12 3 32 9 2 5 12 13 50

50
30 5
5
34(X) 50 27,300 10 127 18 50 27 18 20 9 16 29

<3 <3 <3 <3 <3 <3 <3 13 <3 <3 8 8 15

518 372 456 I 040 1270 450 450 1230 1140 1890 3770

25 88 71 63 673 112 112 113 25 I5 9

43 38 46 160 169 59 59 215 165 326 371



37 145 159 163 1290 284 284 299 35 16 19

41 75 67 168 570 103 103 314 190 367 357

2450 435 8460 505 212 458

4120 68 14.4on 229 16 37

120 24 300
63 19

May 1989 110 < 10 < 10
distance

concentration

2-2 1

1988
CU

Cd COII

downstream

greater

< 10 308 41 I 122 144 110 from

than O.oOl-pm

in micrograms

Climax, filter;

Colorado,

Precision,

in kilometres;

values

are given

Diss.

concentration

in the

less than

for a low and a high concentration.

O.ot)l-pm

based on a linear

9720 <3 177 22 28,400 5700 348 <3 109 <3 40 <3 filtrate: equation;

Coil. all

per liter

tration of the ferrous iron was below detection, a generally oxidizing value for redox potential was used in the geochemical modeling. Temperature and pH were measured in the field. Alkalinity was measured by Gran titration of a 0.45ym filtered, unacidified sample (Stumm and Morgan, 1981). Reagent-grade water was processed through all the equipment h the field to evaluate potential contamination. Values for these process blanks are listed in Table 2 along with the detection limits and analytical precision for each metal, as reported in Friedman and Erdmann (1982). Precision is given as a function of concentration, so Table 2 contains a calculation at the low and high end of concen-

trations measured for the samples. For Cu and Pb the range is from 0.1 to 15pg I-’ because of the lower detection limits. For Fe, Mn, and Zn the range is from 1 to 1,000 pegI-‘. For all the metals, the equipment blank was greater than the detection limit, so the values for the equipment blanks should be considered a lower detection limit, and also a lower limit for precision. Values in Table 2 are reported according to the published detection limits, but can only be used according to these limitations. On both sampling trips, fine-grained bed sediment was collected from depositing areas of the river channel, such as pools, eddies, and bars near the active channel. The assumption was made that fine-grained sediment collected

Metal transport in an acid mine impacted Table 3. Concentration of selected metals in Site

Distance

river

291

suspended colloids in the Arkansas

Al

As

Cd

Cu

Fe

River

Ti

Mn

Pb

Zn

Low flow, October

1988 Gulch River, below Cal. Gulch River, near Malta

25 25 32

55,000 72,000 55,000

250 40 70

120 53 75

1300 150,000 1900 66,000 1200 94,000

1200 1600 1400

2900 2800 7400

3500 loo0 1200

14,000 15,000 19,000

High flow, May 1989 Arkansas River, below Cal. Gulch Arkansas River, at Parkdale Arkansas River. at Portland

25 183 215

79,560 69,400 66,000

210 <20 20

70 I4 13

973 153,000 78 40,100 96 44,400

1500 3400 3300

1840 2280 2590

6730 238 248

16.800 2220 2300

California Arkansas Arkansas

Distance,

in kilometers

downstream

from Climax,

Colorado:

from depositional areas during low flow was representative of the sediment that is transported in suspension during high flow. Sediments were sieved in the field through a 60;um nylon mesh, rinsing with native streamwater. At high flow, in May 1989, suspended sediment also was collected from cableways at gaging stations. Enough solid material for chemical analysis was collected from about 50 I of water using tangential-flow filtration to remove water. Sediment samples were kept on ice during transport. Subsamples were taken for chemical extraction and the remaining sediment was freeze dried for total digestion. Analysis of the less-than-60;um sediment included an extractable fraction and a total digestion. The extractable fraction contained metals associated with Fe and Mn oxides on the surface of grains and the metals in colloids (Chester and Hughes. 1967). The procedure included partial dissolution using I M hydroxylamine hydrochloride in 25 volume percent acetic acid. Metal concentrations in the extracted solutions were determined using flame-atomic absorption for October 1988 samples and ICP-AES for May 1989 samples. The total digestion used a combination of nitric, hydrofluoric, and perchloric acids (Lichte et al., 1987). Metals in the total digestion were analyzed by ICP-AES, and values of precision are reported by Arbogast (1990).

RESULTS

AND DISCUSSION

Ultrafiltration helped to distinguish between the dissolved and colloidal fractions for each of the metals. Metal concentrations of colloids in Table 2 represent the indirect measurements; concentrations in Table 3 represent the direct measurements (see Methods section). Colloidal concentrations in the water indicate partitioning between the dissolved and colloidal fractions and allow the calculation of loads in the river. Concentrations in the solid are a better indication of the chemical composition of the colloids. Determinations of metals in fine-grained bed sediments are listed in Table 4, including both extractable and total metals. Data in each table are grouped by mainstem and tributary samples in downstream order to emphasize the attenuation of metal concentrations in both the dissolved and colloidal fractions. Dissolved and colloidal fractions

Dissolved and colloidal concentrations from Table 2 were multiplied by the discharge in Table 1 to

all concentrations

in parts per million

convert concentrations to loads of metals, in kg day-’ Increases in load generally indicate sources of metal loading and decreases indicate the results of physical, chemical, and biological processes that remove metals from the water. Clear distinctions occurred among the metals with respect to partitioning between dissolved and colloidal fractions. The partitioning changed spatially for some metals during transport downstream and also differed temporally between high and low flow periods. In this study, Fe dominated the colloidal fraction; Fe was consistently greater in the colloidal fraction than in the dissolved fraction (Table 2). Most of the precipitation of colloidal Fe actually occurred in California Gulch, about 1 km upstream from the confluence with the Arkansas River. Samples of suspended colloids from California Gulch (25 km) indicate the composition of the Fe-colloids soon after formation in the stream (Table 3). Arsenic, Cd, Cu, Pb, and Zn occurred in substantial concentrations. rendering the Fe colloids toxic to aquatic organisms (Lee, 1975; Jenne, 1977). If the colloids in the water and in the bed sediment enter the food chain of fish, the colloids could be a pathway leading to chronic toxicity. There are three main sources of colloidal Fe during low flow (Fig. 3). The load increased by 2.74 kg day-’ downstream from the Leadville Drain, 47.7 kg day-’ downstream from California Gulch, and 30.8 kg day-’ downstream from Chalk Creek. A decrease of 29.7 kg day-’ occurred between 25 and 71 km, indicating the sedimentation of aggregated colloids to the streambed. There was not a similar loss downstream from 111 km. This general pattern indicates that colloidal Fe occurs throughout the study reach at all times during the year, and not just during surges of water from mining areas. Colloidal Fe loads were much greater during high flow than low flow (Fig. 3). The three main sources mentioned above contributed greater loads during high flow than during low flow (Table 2). In addition, Tennessee Creek contributed about 175 kg day-’ that likely comes from the flushing of metals from a wetland in St. Kevin Gulch (Walton-Day, 1991). These greater inflows, however, do not account for the total mainstem loads. During low flow, the load

B. A. Kimball et al.

292

Table 4. Concentration of extractable and total metals in bed sediments from the Arkansas River basin

Al

Fe C

(weight percent) Tot

Ext

CU

Ext

Tot

Ext

-

0.19

-

(weight percent) Tot

1.95

Arkansas River, 20 0.08 25 0.03 32 0.07 46 0.06 71 0.06 96 Ill on;4 120 0.04 150 0.05 183 0.03 194 0.05 215 0.05

October 1988 6.7 46 6.65 13 5.25 57 6.05 20 6.25 14 6.5 11 7.1 3.7 6.75 3.3 7.15 3.6 7.15 3.6 6.7 3 6.1 1.1

97

Pb

Ext

Tot

Ext

Tot

Ext

-

4.22

-

937

-

Average Igneous rock* -

Mn

46 74 59 21 14 11 5 3 4 4 3 1.1

39 71 890 1070 56 480 24 121 23 76 14 65 2.3 47 1.7 40
2.6 6 3.5 0.9 1 0.6 0.2 0.1 0.2 0.2 0.3 0.3

4.57 9.03 5.97 4.39 4.39 4.28 4.11 3.55 3.67 3.58 3.93 3.25

4600 2000 1300 1800 2200 1800 670 620 430 740 590 370

4640 2390 1690 1960 2400 2130 1140 1040 841 1150 1020 655

1800 2460 28 62 14 62 34.5 83 3.2 48 0.9 33 1.9 34 2 29 Cl 41 48.2 22

8.8 0.3 1.6 0.78 1 0.3 0.3 0.3 0.3 0.2

17.6 3.47 5.07 5.62 5.25 3.76 4.56 4.26 3.32 2.3

2500 570 1300 235 430 280 1200 270 400 490

3060 2700 997 41 1820 57 960 190

220 1700 1400 450 260 220 100 73 92 62 68 38

Zn Tot 16 222 2830 1560 526 273 241 143 98 111 84 81 49

Ext

9000 22,000 13,000 3100 2300 1900 650 670 740 760 670 250

Tot 80 8280 24,300 12,900 3600 2680 2160 925 708 946 951 846 374

Tributaries, October 1988 25 0.41 4.95 42 0.01 6.3 74 1.3 6.25 86 0.06 6.95 117 5.3 129 0% 5.95 162 0.01 5.8 183 0.03 7.2 202 0.01 6.3 223 ns 5.05

97
97
Arkansas River, 24 0.17 25 0.25 32 0.37 111 0.18 183 0.31 215 0.05

cl
16 26 49 9 7 4


59 359 297 58 52 40

1.3 2.55 4.05 0.54 0.56 0.27

3.55 6.22 5.94 4.43 4.1 4.07

2750 3850 9000 1385 1350 235

3120 4370 9040 1830 1540 934


249 2040 1190 221 45 84

3150 6900 9000 1445 1500 485

3280 7370 9390 1680 1520 863


28 6

770
858 83

13 0.43

16.2 5.47

1550 575

2010 1500


6570 248

9300 570

9.580 1110

May 1989 7.34 7.95 7.48 6.6 7.25 7.27

Tributaries, May 1989 25 0.74 7.12 86 0.77 6.28

Distance, in kilometers downstream from Climax, Colorado; all concentrations From Hem (1985).

downstream

from each tributary was consistently

than or equal to the sum of the upstream

load

less

plus the contribution of the particular tributary, with the exception of California Gulch. During high flow, the mainstem load downstream from tributaries was greater than the sum of the upstream load plus the tributary load. For some sites the gain is more than twice the tributary load. This difference likely indicates that resuspension of colloids all along the mainstem channel is a likely source of colloidal Fe load during runoff. Avery rough comparison can be made between the quantity of Fe deposited on the streambed during low flow and the amount transported during high flow. The hydrograph [Fig. 2(a)] suggests that the flushing likely occurred in April and May, during the rising limb of the hydrograph. Between 25 and 71 km, 29.7 kg day-’ was lost during low flow and during high flow about 1260 kg day-’ was gained in the same

1150 634 1660 647 689 857

25 17 13 12 47 20

7.4 55 72 280 38 23 32 24 62 23

26.000 30,900 44 163 110 210 120 517 130 264 ‘Cl 95 17 136 13 101 260 406 63 107

in parts per million.

reach. A simple ramp model suggests that roughly 75,700 kg of Fe were transported out of the Leadville area during runoff, compared to 8760 kg deposited during low flow. Thus, much more Fe is transported away from Leadville than can be accounted for by loss to the streambed. This suggests that the flushing involves Fe from mines and streambeds throughout the Leadville area and not just from the tributaries that were sampled here. Comparison of Fe concentrations in different filtrates from mainstem samples indicates that concentrations of Fe in the 0.001~pm filtrate were significantly lower than in the 0.45~pm filtrate at low flow and lower than both the O.l- and 0.4.5;um filtrates at high flow (Fig. 4). The difference between the O.OOl-and 0.45pm concentrations was greatest in the Leadville area, near the source of mine drainage. Downstream from there, the colloidal fraction became less dominant, reflecting the settling of some

293

Metal transport in an acid mine impacted river

-DISSOLVED -COLLOIDAL -COLLOIDAL

I.

0.01 0

I..

I

I

I

I

I

100

50

I * 150

HIGH FLOWLOW FLOW HIGH FLOW i

1.. 200

DISTANCE, IN KILOMETERS Fig. 3. Variation of dissolved and colloidal Fe loads during low and high flow sampling.

??

. .

MEAN LOWER NOTCH UPPER NOTCH

Fig. 4. Boxplots comparing concentrations of Fe in various filtrates of water from the Arkansas River. The notches represent a test of significant differences among the groups. Overlapping notches indicate that groups are not significantly different (Velleman and Hoaglin, 1981).

of the colloidal load. Note that the significant difference among groups in Fig. 4 applies to the whole set of data, not just to individual sample sites. This result emphasizes the large difference among the filtrate concentrations. Concentrations of Fe in the O.l- and 0.45;um filtrates during high flow were not significantly different from each other, but both were significantly greater than the O.OOl-pm filtrate con-

centrations (Fig. 4). Without considering the ultrafiltrate, one might erroneously assume that the 0.45 or the 0.1~pm filtered Fe adequately represented dissolved Fe in streams with Fe-rich colloids (LeGendre and Runnells, 1975). The practical effects of this difference lie in calculations for reactive solute transport and also in the evaluation of water with respect to water-quality standards. Toxic metals that are

B. A. Kimball et al.

294

-15

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pH, IN LOG UNITS Fig. 5. (a) Variation of ferric iron activity with pH. The lines indicate linear regression through the two sets of data. (b) Variation of Fe, filtered through different pore-sized membranes, with pH. The line of - 1 slope indicates the trend expected for an equilibrium control by amorphous ferric hydroxide.

actually part of the colloidal fraction mistakenly could be considered as part of the dissolved fraction. Partitioning of ferric iron to the colloidal fraction between pH 7 and 8 is by precipitation according to

the reaction: Fe(OH):

+ HZ0 ti Fe(OH),,,,

+ H+

(I)

Thermodynamic calculations, using either the O.OOlor 0.45;um filtrate Fe concentration indicated that the samples were supersaturated with respect to amorphous Fe(OH),,,, [Fig. 5(a)]. In a plot of ferric ion activity versus pH, a line representing equilib-

rium with amorphous Fe(OH),,,, would have a slope of -3. Data for both filtrates generally indicate that slope. Although this result might indicate that the filter size does not matter in determining equilibrium, the procedure of calculating activities as a function of pH and then regressing against pH introduces a forced correlation between the ferric ion activity and pH. Neal et al. (1987) have shown that the same -3 slope can result from using random numbers as input for Fe concentrations and then calculating speciation. Thus, the calculation of ferric iron activity and its correlation to pH is not a sufficient test for equilib-

295

Metal transport in an acid mine impacted river

i 0.11

,‘I

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’

’

’

’

’

’

’

’

’

’

’

100

50

’

LOW FLOW HIGH FLOW LOW FLOW HIGH FLOW

’

’

’

’

’ ’

150

200

DISTANCE, IN KILOMETERS Fig. 6. Variation of dissolved and colloidal Zn loads during low and high flow sampling.

rium. A better indication of equilibrium is the systematic decrease of analytically determined dissolved Fe with the increase in pH. Values of Fe measured in the O.OOl-, O.l-, and 0.45;um filtrates are plotted against pH in Fig. 5(b). A line with a slope of -1 represents equilibrium with amorphous Fe(OH& between pH 7 and 8 [Eq. (l)*], and is drawn near the data points for the 0.001~pm filtrate. The O.OOl;um data do indicate a decrease with pH, but the 0. l- and 0.45pm data do not. Thus, the indication of equilibrium can be obtained using ultrafiltration. Dissolved metals fall into two groups based on the magnitude of instream concentrations; dissolved Mn and Zn had relatively high concentrations, and Al, Cd, Cu, Fe, and Pb had relatively low concentrations. Loads of dissolved and colloidal Zn illustrate the general pattern of metal loading for Zn and Mn (Fig. 6). Dissolved Zn load increased about 18 kg day-’ downstream from the inflow of the Leadville Drain, 28 kg day-’ from the inflow of California Gulch, 14.3 kg day-’ downstream from Lake Fork (26 km), and 49 kg day-’ downstream from Lake Creek (42 km). Some of the increase between 25 and 32 km also could have resulted from the inflow of metal-rich seeps along the stream channel (Patterson, 1988, University of Colorado, pers. commun.). The seeps obtain metals from mine tailings that have been dispersed in the alluvial material downstream from the mining areas. At low flow, the dissolved fraction predominated over most of the study reach, but colloidal Zn also occurred, and persisted downstream to 111 km. The dissolved Zn load decreased *In the pH range of these waters, Fe(OHg would be the predominant Fe species. Thus, the field data follow a slope of - 1 rather than a slope of -3 that results from using Fe3+ as the dominant species.

between 46 and 111 km. Part of this decrease can be accounted for by the concurrent increase in colloidal Zn load, but much of the mass was lost to the bed. An additional increase of about 39 kg day-’ occurred between 111 and 120 km, but the inflow from the South Fork of the Arkansas River accounted for less than 1 kg day-‘. The additional Zn load may be from drainage of old mining or smelting operations near Salida. The dissolved Zn load decreased steadily downstream from 120 km, indicating that the Zn partitioned to the colloidal fraction and to solids in the bed sediment. Loads of both dissolved and colloidal Zn were very different between high and low flow (Fig. 6). During high flow, colloidal Zn predominated over dissolved Zn in most of the study reach and the total Zn load was much greater than during low flow. The difference between high and low flow may be attributed to resuspension of aggregated colloidal material from the streambed; the pattern is similar to colloidal Fe. The patterns of colloidal Fe and Zn both suggest an annual pattern. During low flow, dissolved and colloidal loads decrease downstream as metals partition to the colloidal fraction and the aggregated Fe colloids settle to the streambed. These colloids are resuspended during high flow; there is a flushing of metals with snowmelt runoff creating the greatest loads during the year. There could also be flushing during thunderstorm runoff. Horowitz et al. (1990) observed a doubling of Cu and Zn concentrations in the suspended sediment while sampling during a thunderstorm at 215 km in 1988. Partitioning of Mn was similar to Zn. The greatest dissolved load occurred at 25 km, downstream from the inflow of California Gulch (Table 2). At low flow Mn was mostly in the dissolved fraction. This result is consistent with field data (Laxen and Chandler, 1983)

296

B. A. Kimball etal.

and thermodynamic predictions (Sung and Morgan, 1981; Hem, 1985) for similar conditions of pH and redox in rivers. Dissolved Mn load decreased between 32 and 96 km, corresponding to an increase of Mn in the colloidal load and in the bed sediment (Tables 3 and 4). Between 25 and 32 km, the Mn concentration in the colloidal fraction increased from 2800 to 7400 ppm, similar to the increase of Zn. During high flow colloidal Mn predominated, suggesting resuspension of aggregated colloids from the streambed. Thus the presence of colloids substantially influenced the partitioning of both Mn and Zn and the seasonal nature of the loads for these metals. Although dissolved loads of Cd, Cu, and Pb were not as high as loads of Mn and Zn, the influence of inflows from the Leadville area was clear (Table 2). Generally, mainstem loads of dissolved Cd, Cu, and Pb were highest in the area just downstream from California Gulch and remained essentially constant downstream from the Leadville area. Loads of colloidal Cd, Cu, and Pb were substantial at 25 and 32 km, but decreased below detection downstream from 32 km. Because concentrations of these metals were relatively low, the downstream patterns are not as clear as the patterns for Mn and Zn and the error in calculating loads by multiplying large discharges by small differences in concentration is considerable. Nevertheless, the dissolved and colloidal concentrations of these metals were mostly above detection. The occurrence of Cu, Cd, Pb, and Zn in the dissolved and colloidal fractions can impact aquatic organisms. Differences in dissolved metal concentrations between high and low flow are greatest for Fe. The concentrations of Fe were greater during high flow than during low flow (Fig. 4). The difference could be a response to pH. The 0.001~pm Fe concentration generally follows the slope of -1 [Fig. 5(b)], and at lower pH the equilibrium concentration would be higher. Concentrations of Cd, Cu, Mn, Pb, and Zn were not significantly different between high and low flow periods. The equal concentrations between high and low flow indicate that the high flow concentrations are not diluted; equal concentrations result in much greater loads during high flow than low flow. Concentrations of dissolved and colloidal Al differed from the other metals. Concentrations of Al in most samples were below detection (about 1Opg I-‘) with the exception of samples collected during high flow. These samples indicated detectable Al in the colloidal fraction. At near-neutral pH, dissolved Al concentrations in equilibrium with gibbsite or any aluminosilicate phase should be low (Hem, 1985; Drever, 1988; May, 1992). The absence of Al in the dissolved fraction defined by ultrafiltration supports the thermodynamic predictions of very low concentrations. The detection of dissolved Al concentrations in samples from other rivers affected by

acidic mine drainage, but with near-neutral pH could be a result of colloidal Al passing through 0.45pm filters (e.g. Davis et al., 1991). However, there could also be environmental factors, perhaps complexation by natural organic ligands could allow detectable concentrations of Al to occur in other streams.

Bed sediments

There are four characteristics of metals in bed sediments that appear to distinguish the effects of mining activities. These characteristics are illustrated by Zn concentrations in Fig. 7. First, as noted by Axtmann and Luoma (1989), there is a characteristic pattern of attenuation downstream from the source of the metals [Fig. 7(a)]. Second, the concentrations of metals in mainstem bed sediments generally are greater then average crustal rock abundance that is indicated in Table 4 and also are greater than the concentration in tributary bed sediments. Third, most of the metal concentration is extractable, indicating that the metal is either part of mineral surface coatings or part of aggregated colloids that have settled to the bed. Fourth, the ratios of metals from mine drainage remain constant downstream from a given source of metal contamination (Moore, 1993). For example, the ratio of other metals to Zn in bed sediments should be a characteristic that can distinguish the signature of metal sources, like California Gulch, on the bed sediments downstream. If the metal to Zn ratio in bed sediments affected by California Gulch is preserved in sediments downstream, then inflows of sediment from unaffected tributaries downstream only dilute that ratio and the chemical signature from the Leadville area remains evident. On the other hand, if the ratio of metals to Zn in mainstem sediments changes toward the ratio in tributary bed sediments, then mixing with the tributary sediments masks the effects of mining on bedsediment chemistry. Based on these characteristics, there are four significantly different patterns among the metals in this study. Zinc is an appropriate index metal for comparison to other metals in bed sediments because of its substantial concentrations in the bed sediment and colloidal material (Table 2) and the large percentage of extractable Zn in bed sediments (Table 4). Partitioning of Zn from the water to colloids and bed sediment is consistent with the decrease in dissolved Zn load from the water downstream from 46 km (Fig. 6). Partitioning could be due to sorption of Zn onto the Fe colloids, to coprecipitation, or to precipitation of a Zn mineral phase. Precipitation of Zn may occur according to the reaction: Zn*+ + 2H 2 0 --

Zn(OW2(amorp)

+

a-~+

(2)

The reaction results in a slope of -2 if the log of Zn” activity is plotted against pH. Thermodynamic calculations indicate an appropriate slope for the reaction

291

Metal transport in an acid mine impacted river

,

-S-

a

t

1,000

RIVER, TOTAL 0 TRIBUTARY, TOTAL -O-RIVER, EXTRACTABLE 0 A A

,-

a

TRIBUTARY. EXTRACTABLE CALIFORNIA GULCH COLLOIDS MAINSTEM COLLOIDS A

.II-------e------e

m-m-m

A

-a

0

0

tl”“““““““‘3”“‘I 0

50 ,

o -

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100 I

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DISTANCE, IN KILOMETERS I I, I1 I I, I1 11,

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200 I

I

I

I

I

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Ir

??

LOW FLOW 0 HIGH FLOW -EQUILIBRIUM

WITH Zn(OH),,

??

pH, IN LOG UNITS

Fig. 7. (a) Variation of extractable and total Zn in bed sediments and colloids with downstream distance in October 1988. (b) Activity diagram of dissolved Zn versus pH for samples from the Arkansas River mainstem.

[Fig. 7(b)]. Unlike Fe species, the aqueous Zn hydroxide species are not dominant in solution, and a trend of Zn2+ activity with pH is not a forced correlation; it is a suggestion of equilibrium. The calculated Zn2+ activities, however, were consistently about two-orders of magnitude too low for equilibrium with the amorphous Zn(OH)2t,) phase in the WATEQ4F data base. This could indicate a less amorphous phase is responsible for this control on Zn*+ activities or else that sorption, rather than precipitation, is the dominant control on Zn concentrations. Although the stoichiometry of sorption also produces a slope of -2 for the log of Zn2’ activity versus pH, the high concentration of Zn in the bed sediment is more consistent with precipitation of Zn

as a surface coating. These field data alone cannot clearly distinguish precipitation from sorption. Concentrations of Pb [Fig. 8(a)] and Cd [Fig. 9(a)] follow the same pattern as Zn that is suggestive of mining influences. The majority of Pb and Cd in bed sediments was extractable, which is consistent with sorption or precipitation of Pb and Cd onto colloids and bed sediments. Mainstem concentrations in bed sorptions were elevated by the inflow of California Gulch, and then decreased downstream. These concentrations were consistently higher than tributary concentrations along the entire study reach with the exception of Chalk Creek [86 km, Fig. 8(a)]. The concentrations were also consistently higher than average igneous rock concentrations of these metals

B. A. Kimball eral.

298

*,

,

,a,,

: (4

,

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,

,

,

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TRIBUTARY, EXTRACTABLE TRIBUTARY, TOTAL CALIFORNIA GULCH COLLOIDS MAINSTEM COLLOIDS

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i

1;2fg3 . . . . .’. . . 223 . . . . .._._....................................................

10

n liE’ 100

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ZINC, IN PARTS PER MILLION

Fig. 8. (a) Variation of extractable and total Pb in bed sediments and colloids with downstream distance in October 1988. (b) Variation of Pb with Zn in bed sediments and colloids from the Arkansas River, October 1988 and May 1989. Line of 1:l slope indicates dilution of concentrations at a constant Pb to Zn ratio.

that are indicated by the vertical and horizontal lines in Figs 8 and 9. The concentrations of Pb or Cd in tributary bed sediments was not sufficient to affect the ratios of either metal to Zn; the ratio from California Gulch occurred all the way downstream [Figs 8(b) and 9(b)]. Figures 8(b) and 9(b) represent mixing of three main sources that combine to influence the bedsediment chemistry. The first source is the input of

sediment upstream from California Gulch; this is represented by the sample at 20 km. The second source is the input of sediment and colloids from California Gulch at 25 km. The third source is the input from the tributaries downstream from California Gulch. Among these downstream tributaries, Badger Creek (129 km), Tallahassee Creek (182 km), and Fourmile Creek (202 km) are the major contributors of sediment. Bed sediments from these

299

Metal transport in an acid mine impacted river

I

I

I

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I

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25

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TRIBUTARIES. LOW FLOW RIVER, HIGH FLOW TRIBUTARIES, HIGH FLOW CALIFORNIA GULCH COLLOIDS MAINSTEM COLLOIDS

a

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I I I1111

I

I

I In.111

I

I

I

10,000

1,000 ZINC. IN PARTS PER MILLION

Fig. 9. (a) Variation of extractable and total Cd in bed sediments and colloids with distance in October,

1988. (b) Variation of Cd with Zn in bed sediments and colloids from the Arkansas River, October 1988 and May 1989. Line of 1:l slope indicates dilution of concentrations at a constant Cd to Zn ratio.

tributaries had relatively low concentrations of metals compared to the tributaries influenced by mining (Table 4). Mixing of upstream sediments with sediments from California Gulch resulted in bedsediment concentrations at 25 km that fall along a dashed line representing mixing between the sample from California Gulch (25 km) and the mainstem site at 20 km [Fig. 8(b)]. At 25 km, the mixture would physically consist of fine-grained sediment from

upstream and the colloids and fine-grained sediment from California Gulch. Tributary inflows downstream from California Gulch dilute the total concentrations and do not change the metal Zn ratios from 25 km [Fig. g(b)]. The second pattern of bed-sediment concentration is indicated by the spatial distributions of Fe and Cu. The pattern of Fe indicates that California Gulch (25 km) was a principal source of total and extractable Fe

300

B. A. Kimball et al.

-m-RIVER, 8 100,000

:

A b H 11 0

>-#?._Jm

TOTAL

0 0

TRIBUTARY, EXTRACTABLE TRIBUTARY. TOTAL

A

CALIFORNIA GULCH COLLOIDS

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TRIBUTARIES, HIGH FLOW CALIFORNIA GULCH COLLOIDS MAINSTEM COLLOIDS

........

10,000 IO

1,000

100

10,000

ZINC, IN PARTS PER MILLION

Fig. 10. (a) Variation of extractable and total Fe in bed sediments and colloids with distance in October 1988. (b) Variation of Fe with Zn in bed sediments and colloids from the Arkansas River, October 1988 and May 1989. Line of 1:l slope indicates dilution of concentrations at a constant Fe to Zn ratio.

in bed sediments [Fig. 10(a)]. The relatively high concentrations of Fe in the bed sediments near the Leadville area likely result from settling of the Fe colloids from the water to the streambed in the first 50 km downstream from California Gulch; colloidal and bed-sediment concentrations are nearly equal [Fig. 10(a)]. This corresponds to the decrease of colloidal Fe load between 32 and 71 km (Fig. 3). The sample from 32 km indicates the same Fe to Zn ratio as the sample from 25 km. At 46 km, however, there is a

definite shift away from the dilution line; an increase in the Fe to Zn ratio occurs with mixing. From 46 to 71 km and then to 111 km, the ratio changes toward the ratios in the tributary sediments, which is similar to the Fe to Zn ratio in average crustal rocks [where the dashed lines cross, Fig. 10(b)]. Thus, Fe concentrations in bed sediments near Leadville were dominated by colloids, but downstream from the mining areas mixing with tributary sediments alters that chemical signature for Fe in bed sediments.

Metal transport in an acid mine impacted river

301

C”“I”“I”“I”“l”‘~

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RIVER. EXTRACTABLE RIVER. TOTAL TRIBUiARY, EXTRACTABLE TRIBUTARY, TOTAL CALIFORNIA GULCH COLLOIDS MAINSTEM COLLOIDS

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Fig. 11. (a) Variation of extractable and total Cu in bed sediments and colloids with distance in October 1988. (b) Variation of Cu with Zn in bed sediments and colloids from the Arkansas River, October 1988 and May 1989. Line of 1:l slope indicates dilution of concentrations at a constant Cu to Zn ratio.

The distribution of extractable and total Cu in bed sediments was comparable to the distribution of Fe in fine-grained bed sediments [Fig. 11(a)], but the pattern of the Cu to Zn ratio differs from the Fe to Zn ratio. River bed sediments from 25 to 96 km had Cu to Zn ratios that indicate the signal of California Gulch. The Cu concentration in these samples was higher than the downstream samples. Samples downstream from 46 km, however, had Cu concentrations near the average igneous rock value [dashed horizon-

tal line in Fig. 11(b)], and they are distinguishable from tributary inputs only by their higher Zn concentrations. Thus, dilution by tributary sediments controls the concentrations of Cu in mainstem sediments downstream from 46 km. The precipitation of Cu*’ in an oxide or hydroxide may yield a slope of -2 in a plot of cu*+ activity versus pH, but the thermodynamic calculations did not indicate a consistent slope. A third pattern of metal concentrations in bed

B. A. Kimball er al. I

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DISTANCE, IN KM

__ ___

RIVER, LOW FLOW TRIBUTARIES, LOW FLOW RIVER, HIGH FLOW TRIBUTARIES. HIGH FLOW CALIFORNIA GULCH COLLOIDS MAINSTEM COLLOIDS

100

20 .** , 24.8

1,000

:

24.9

10,000

ZINC, IN PARTS PER MILLION

Fig. 12. (a) Variation of extractable and total Mn in bed sediments and colloids with distance in October 1988. (b) Variation of Mn with Zn in bed sediments and colloids from the Arkansas River, October 1988 and May 1989. Line of 1:l slope indicates dilution of concentrations at a constant Mn to Zn ratio.

sediments is seen in the concentration profile of Mn. The highest concentrations of Mn occurred upstream, rather than downstream from California Gulch. This likely indicates precipitation of Mn from the Leadville Drain onto bed sediments [Fig. 12(a)]. The concentration in bed sediments increased between 32 and 71 km, and the additional Mn was extractable, indicating it was on the surfaces of bed sediments. This caused a substantial shift in the ratio

of Mn to Zn in the bed sediments [Fig. 12(b)]. The colloidal fraction and the river sediments from 25 and 32 km during low flow had the same ratio as the colloids and the bed sediments from California Gulch, but precipitation or sorption of Mn caused all the other river samples to have a greater ratio. The higher ratio also represents the ratio of Mn and Zn from the Leadville Drain inflow (15 km, no sediment or colloid sample). During high flow, all the river bed

303

Metal transport in an acid mine impacted river

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Fig. 13. (a) Variation of extractable and total Al in bed sediments and colloids with distance in October 1988. (b) Variation of Al with Zn in bed sediments and colloids from the Arkansas River, October 1988 and May 1989. Line of 1:l slope indicates dilution of concentrations at a constant Al to Zn ratio.

sediments reflected this higher ratio of Mn to Zn, even the samples from 25 and 32 km. Thus, Mn is reactive, and the relatively high concentrations of dissolved Mn that come from California Gulch partition to the sediments and colloids downstream. The distinct pattern for the Mn to Zn ratio is likely a result of the reactive nature of Mn rather than a result of mixing like the other metals. Concentrations of Al in bed sediment followed a fourth pattern which was very different from all the

other metals [Fig. 13(a)]” There was only a small percentage of extractable Al in bed sediments, suggesting that Al mostly occurred in silt- and claysize aluminosilicate minerals. Unlike other metals that mostly were contributed by mining, total Al concentrations in bed sediments are dominated by tributary inflows, and the Al to Zn ratio steadily increased toward values of tributary bed sediments and generally toward the ratio of average igneous rocks [Fig. 13(b)].

B. A. Kimball et al.

304

Samples of colloids from 183 and 215 km are plotted in Figs 7-13. Concentrations of Cd, Cu, Mn, Pb, and Zn in these colloid samples are equal to bedsediment concentrations in samples from 46 to 96 km, but are not equal to bed-sediment concentrations at their respective distances downstream. These colloidal samples also have metal to Zn ratios that fall on the 1: 1 dilution lines. Both these observations suggest that the colloids are actually formed upstream and kept in suspension. Because the colloids in suspension do not mix with tributary bed sediments, they retain the signature of the mining areas upstream.

port of metals may help with decisions for remediation of their effects. Acknowledgements-Support for this work was from the Toxic Substances Hydrology Program of the U.S. Geological Survey, Water Resources Division. Patrick Edelmann, Bryan Nordland, and Gregory Wetherbee assisted with sample collection and analysis. The manuscript benefited from the review of Howard Taylor and Robert Stallard of the U.S. Geological Survey, Donald Runnells, and an anonymous reviewer. Editorial handling: Dr Don Runnells.

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