textile industries release large amounts of chromium to surface waters (Fishbein 1981). Disposal of

chromium-containing commercial products and coal ash from electric utilities and other industries are the

major sources of chromium release into the soil (Nriagu and Pacyna 1988). Solid waste and slag

produced during the roasting and leaching processes of chromate manufacturing when disposed of

improperly in landfill or when used as fill can be potential sources of chromium exposure.

Chromium released into the environment from combustion processes and ore processing industries is

from coal-fired power plants (Stern et al. 1984) and from chromate manufacturing and user sites.

Chromium is primarily removed from the atmosphere by fallout and precipitation. By analogy with

copper, the residence time of chromium in the atmosphere is expected to be <10 days (Nriagu 1979).

Most of the chromium in lakes and rivers will ultimately be deposited in the sediments. Chromium in the

CHROMIUM 282

5. POTENTIAL FOR HUMAN EXPOSURE

1-9

10-18

19-30

31-44

52-67

91

Frequency of

NPL Sites

Derived from HazDat 2000

CHROMIUM 283

5. POTENTIAL FOR HUMAN EXPOSURE

1

2

3

4-5

10-12

16

Frequency of

NPL Sites

Derived from HazDat 2000

CHROMIUM 284

5. POTENTIAL FOR HUMAN EXPOSURE

aquatic phase occurs in the soluble state or as suspended solids adsorbed onto clayish materials, organics,

or iron oxides. Most of the soluble chromium is present as chromium(VI) or as soluble chromium(III)

complexes and generally accounts for a small percentage of the total. Soluble chromium(VI) may persist

in some bodies of water for a long time, but will eventually be reduced to chromium(III) by organic

matters or other reducing agents in water (Cary 1982; EPA 1984a). The residence times of chromium

(total) in lake water range from 4.6 to 18 years (Schmidt and Andren 1984). Chromium(III) in soil is

mostly present as insoluble carbonate and oxide of chromium(III); therefore, it will not be mobile in soil.

The solubility of chromium(III) in soil and its mobility may increase due to the formation of soluble

complexes with organic matters in soil. A lower soil pH may facilitate complexation. Chromium has a

low mobility for translocation from roots to the aboveground parts of plants (Calder 1988; Cary 1982;

EPA 1984a, 1985a; King 1988; Stackhouse and Benson 1989).

The arithmetic mean concentrations of total chromium in the ambient air in United States, urban,

1990b). The chromium concentrations in U.S. river waters usually range from <1 to 30 µg/L, with a

median value of 10 µg/L (Eckel and Jacob 1988; EPA 1984a; Malm et al. 1988; Ramelow et al. 1987;

Smith et al. 1987). The total chromium concentrations in U.S. drinking water range from 0.4 to 8.0 µg/L,

with a mean value of 1.8 µg/L (Greathouse and Craun 1978). In ocean water, the mean chromium

concentration is 0.3 µg/L (Cary 1982). Total chromium concentrations in conterminous U.S. soils range

from 1.0 to 2,000 mg/kg, with a mean of 37.0 mg/kg (USGS 1984). The typical chromium levels in most

fresh foods are <50 µg/kg (Fishbein 1984).

The general population is exposed to chromium by eating food or food supplements, drinking water, and

inhaling air that contain chromium. The mean daily dietary intake of chromium from air, water, and food

is estimated to be <0.2–0.4, 2.0, and 60 µg, respectively (see Section 5.5). Dermal exposure to chromium

may also occur during the use of consumer products that contain chromium, such as wood treated with

copper dichromate or chromated copper arsenate and leather tanned with chromic sulfate. Exposure to

chromium for occupational groups can be two orders of magnitude higher than the exposure to the

general population (Hemminki and Vainio 1984). Occupational exposure to chromium occurs mainly

from chromate production, stainless steel production and welding, chrome plating, production of

ferrochrome alloys, chrome pigment production and user industries, and from working in tanning

industries (Stern 1982). Of the general population, people who reside in the vicinity of chromium waste

disposal sites and chromium manufacturing and processing plants have a greater probability of elevated

chromium exposure.

CHROMIUM 285

5. POTENTIAL FOR HUMAN EXPOSURE

According to the Toxics Release Inventory (TRI), in 1997 a total of 32,811,382 pounds (14,883,243 kg)

of chromium was released to the environment from 3,391 large processing facilities (TRI97 1999).

Table 5-1 lists estimated amounts released from specific anthropogenic sources during an unspecified

time prior to 1990, including the percentage which were in the hexavalent state (EPA 1990b). In

addition, an estimated 272,732 pounds (123,711 kg) were released by manufacturing and processing

facilities to publicly owned treatment works (POTWs) and an estimated 5,788,705 pounds

(2,625,757 kg) were transferred offsite (TRI97 1999). The TRI data should be used with caution because

only certain types of facilities are required to report. This is not an exhaustive list.

According to the Toxics Release Inventory, in 1997 the estimated releases of chromium of

706,204 pounds (320,334 kg) to the air from 3,391 large processing facilities accounted for about 2.2% of

total environmental releases (TRI97 1999). Table 5-2 lists amounts released from facilities. The TRI

data should be used with caution, however, since only certain types of facilities are required to report.

This is not an exhaustive list.

Total chromium has been identified in 45 air samples collected from 1,036 NPL hazardous waste sites

where it was detected in some environmental media (HazDat 2000). Chromium(VI) has been identified

in 4 air samples collected from 120 NPL hazardous waste sites where it was detected in some

environmental media (HazDat 2000).

Continental dust flux is the main natural source of chromium in the atmosphere; volcanic dust and gas

flux are minor natural sources of chromium in the atmosphere (Fishbein 1981). Chromium is released

into the atmosphere mainly by anthropogenic stationary point sources, including industrial, commercial,

and residential fuel combustion, via the combustion of natural gas, oil, and coal. Other important

anthropogenic stationary point sources of chromium emission to the atmosphere are metal industries. It

anthropogenic sources in 1970 (EPA 1984a). These older estimates indicated that emissions from the

metal industry ranged from 35% to 86% of the total, and emissions from fuel combustion ranged from

11% to 65% of the total (EPA 1978). More recent estimates of atmospheric chromium emission in 1976

and 1980 in the Los Angeles, California, and Houston, Texas, areas indicate that emissions from

CHROMIUM 286

5. POTENTIAL FOR HUMAN EXPOSURE

Source category

Estimated

number of

sources

Chromium

emissions

(metric tons/year)

Estimated

hexavalent

chromium (%)

Combustion of coal and oil Many 1,723 0.2

Chromium chemical manufacturing 2 18 67

Chemical manufacturing cooling towers 2,039 43 100

Petroleum refining cooling towers 475 32 100

Speciality/steel production 18 103 2.2

Primary metal cooling towers 224 8 100

Comfort cooling towers 38,000 7.2–206 100

Texile manufacturing cooling tower 51 0.1 100

Refractory production 10 24 1.3

Ferrochromium production 2 16 5.4

Sewage sludge incineration 133 13 <0.1

Tobacco cooling towers 16 0.2 100

Utility industry cooling towers 6 1.0 100

Chrome ore refining 6 4.8 <0.1

Tire and rubber cooling towers 40 0.2 100

Glass manufacturing cooling towers 3 0.01 100

Cement production 145 3 0.2

Municipal refuse incineration 95 2.5 0.3

CHROMIUM 287

5. POTENTIAL FOR HUMAN EXPOSURE

Number of

Underground

injection

Total

transfer

Off-site waste

transfer

AK 1 0 0 0 0 0 0 0

AL 87 24,735 4,043 223,481 0 252,259 907 29,086

AR 49 10,282 557 5,065 0 15,904 1,177 10,450

AZ 31 2,787 0 410,344 0 413,131 653 1,118

CA 147 8,469 1,183 141,237 0 150,889 5,680 31,462

CO 33 2,907 6 1,075 0 3,988 348 10,659

CT 66 5,573 1,353 5 0 6,931 1,213 39,106

DE 6 3,593 0 2,886 0 6,479 0 0

FL 45 3,035 255 2,790 0 6,080 578 32

GA 90 3,969 10,163 212,648 0 226,780 3,642 125,753

HI 2 0 5 0 0 5 0 0

IA 55 25,163 766 2,747 0 28,676 4,573 2,390

ID 6 2,807 0 733,783 0 736,590 0 0

IL 212 38,165 2,292 1,793,553 0 1,834,010 11,162 666,573

IN 199 68,158 6,341 324,246 2,100 400,845 6,095 22,983

KS 38 17,535 250 108,525 250 126,560 1,381 90,430

KY 72 7,575 721 97,804 0 106100 1,185 58,785

LA 34 17,722 1,129 751 0 19,602 7 5,888

CHROMIUM 288

5. POTENTIAL FOR HUMAN EXPOSURE

Number of

Underground

injection

Total

transfer

Off-site waste

transfer

MA 79 3,664 30 1,887 0 5,581 2,267 102,042

MD 32 1,836 1,602 110,605 0 114,043 1,246 23,558

ME 18 910 761 5,217 0 6,888 68,452 202

MI 162 28,689 673 52,956 11,756 94,074 13,373 222,886

MN 49 5,117 24 32 0 5,173 32,210 128

MO 74 28,200 371 10,872 0 39,443 2,285 10,697

MS 49 4,075 1,947 19,219 580,000 605,241 775 18,324

MT 4 505 0 0 0 505 0 0

NC 91 9,630 1,710 9,111,241 0 9,122,581 8,709 19,803

ND 5 97 10 0 0 107 3 0

NE 22 2,094 784 2,505 0 5,383 6,370 38,380

NH 20 5,673 0 0 0 5673 32 1,310

NJ 48 12,069 192 3,600 0 15,861 8,351 36,559

NM 7 1,049 77 330,037 0 331,163 16 5

NV 6 1,282 0 778 0 2,060 0 0

NY 94 18,768 1,810 999,639 0 1,020,217 9,208 40,222

OH 333 118,231 13,385 904,571 250 1036437 9,594 349,302

OK 67 18,491 57 34,788 0 53,336 2,261 5,493

Number of

Underground

injection

Total

transfer

Off-site waste

transfer

CHROMIUM 289

5. POTENTIAL FOR HUMAN EXPOSURE

OR 37 6,346 203 13,397 0 19,946 389 14,693

PA 316 10,1742 10,577 156,594 0 268,913 9,383 227,716

PR 4 69 0 0 0 69 0 0

RI 14 18 23 0 0 41 256 250

SC 72 7,823 919 103,254 0 111,996 2,405 20,552

SD 8 80 0 0 0 80 15 250

TN 81 16,101 1,903 79,383 530,000 627,387 1,317 34,512

TX 177 34,582 1,4201 14,598,106 7,203 14,654,092 26,146 3,204,777

UT 28 3,970 250 230,352 0 234,572 1,285 4,698

VA 51 2,007 1,737 2,100 0 5,844 2,667 22,648

VT 4 0 0 4 0 4 255 0

WA 39 4,623 14,426 1,923 0 20,972 389 20,857

WI 197 23,047 1,566 9,993 0 34,606 24,281 158,314

WV 27 2,941 13,082 18,242 0 34,265 191 115,812

WY 3 0 0 0 0 0 0 0

Source: TRI 97 1999

CHROMIUM 290

5. POTENTIAL FOR HUMAN EXPOSURE

26 to 45% of the total (Cass and McRae 1986). The primary stationary nonpoint source of chromium

emission into the atmosphere is fugitive emissions from road dusts. Other potentially small sources of

atmospheric chromium emission are cement-producing plants (cement contains chromium), the wearing

down of asbestos brake linings that contain chromium, incineration of municipal refuse and sewage

sludge, and emission from chromium-based automotive catalytic converters. Emissions from cooling

towers that previously used chromate chemicals as rust inhibitors are also atmospheric sources of

chromium (EPA 1984b; Fishbein 1981).

According to the Toxics Release Inventory, in 1997 the estimated releases of chromium of

111,384 pounds (50,524 kg) to water from 3,391 large processing facilities accounted for about 0.3% of

total environmental releases (TRI97 1999). Table 5-2 lists amounts released from facilities. The TRI

data should be used with caution, however, since only certain types of facilities are required to report.

This is not an exhaustive list.

Total chromium has been identified in 400 surface water and 1,178 groundwater samples collected from

1,036 NPL hazardous waste sites where it was detected in some environmental media (HazDat 2000).

Chromium(VI) was identified in 32 surface water and 113 groundwater samples collected from 120

hazardous waste sites where it was detected in some environmental media (HazDat 2000).

The most significant anthropogenic point sources of chromium in surface waters and groundwaters are the

waste waters from electroplating operations, leather tanning industries, and textile manufacturing. In

addition, deposition of airborne chromium is also a significant nonpoint source of chromium in surface

water (Fishbein 1981). In a 1972 survey, the contribution of different sources to chromium load in the

influent waste water of a treatment plant in New York City was estimated to be as follows: electroplating

industry, 43%; residential waste water, 28%; other industries, 9%; runoff, 9%; and unknown, 11%

(Klein et al. 1974). On a worldwide basis, the major chromium source in aquatic ecosystems is domestic

waste water effluents (32.2% of the total). The other major sources are metal manufacturing (25.6%),

ocean dumping of sewage (13.2%), chemical manufacturing (9.3%), smelting and refining of nonferrous

metals (8.1%), and atmospheric fallout (6.4%) (Nriagu and Pacyna 1988). Annual anthropogenic input of

chromium into water has been estimated to exceed anthropogenic input into the atmosphere (Nriagu and

CHROMIUM 291

5. POTENTIAL FOR HUMAN EXPOSURE

Pacyna 1988). However, land erosion, a natural source of chromium in water, was not included in the

Nriagu and Pacyna (1988) estimation of chromium contributions to the aquatic environment.

According to the Toxics Release Inventory, in 1997, the estimated releases of chromium of

30,862,235 pounds (13,999,110 kg) to soil from 3,391 large processing facilities accounted for about

94.1% of total environmental releases (TRI97 1999). An additional 1,131,559 pounds (513,275 kg) of

chromium, amounting to 3.4% of the total environmental release, was injected underground. Table 5-2

lists the amounts released from facilities. The TRI data should be used with caution, however, since only

certain types of facilities are required to report. This is not an exhaustive list.

Total chromium has been identified in 939 soil and 472 sediment samples collected from 1,036 NPL

hazardous waste sites where it was detected in some environmental media (HazDat 2000).

Chromium(VI) has been identified in 59 soil and 22 sediment samples collected from 120 NPL hazardous

waste sites where it was detected in some environmental media (HazDat 2000).

On a worldwide basis, the disposal of commercial products that contain chromium may be the largest

significant sources of chromium release into soil include the disposal of coal fly ash and bottom fly ash

from electric utilities and other industries (33.1%), agricultural and food wastes (5.3%), animal wastes

(3.9%), and atmospheric fallout (2.4%) (Nriagu and Pacyna 1988). Solid wastes from metal

manufacturing constituted <0.2% to the overall chromium release in soil. However, the amount of

chromium in sludge or residue that is disposed of in landfills by manufacturing and user industries that

treat chromate wastes in ponds and lagoons is not included in the estimation by Nriagu and Pacyna

(1988).

Chromium is present in the atmosphere primarily in particulate form. Naturally occurring gaseous forms

of chromium are rare (Cary 1982). The transport and partitioning of particulate matter in the atmosphere

depend largely on particle size and density. Atmospheric particulate matter is deposited on land and

CHROMIUM 292

5. POTENTIAL FOR HUMAN EXPOSURE

water via wet and dry deposition. In the case of chromium, the mass median diameter of the ambient

velocity is 0.5 cm/second (Schroeder et al. 1987). This size and deposition velocity favor dry deposition

by inertial impaction (Schroeder et al. 1987). Wet removal of particulate chromium also occurs by

rainout within a cloud, and washout below a cloud, and acid rain may facilitate removal of acid-soluble

chromium compounds from the atmosphere. The wet scavenging ratio (i.e., the concentration of

contaminant in precipitation over the concentration in unscavenged air) ranges from 150 to 290 for

chromium (Dasch and Wolff 1989; Schroeder et al. 1987). The wet deposition ratio increases with

particle size and decreases with precipitation intensity (Schroeder et al. 1987). Chromium particles of

aerodynamic diameter <20 µm may remain airborne for longer periods of time and be transported for

greater distances than larger particles. The monthly dry deposition flux rate of chromium measured in

values occurring during the winter months (Morselli et al. 1999).

A maximum of 47% of the total chromium in ferrochrome smelter dust may be bioavailable as indicated

by acid/base extraction. About 40% of the bioavailable chromium may exist as chromium(VI), mostly in

chromium particles are transported from the troposphere to the stratosphere (Pacyna and Ottar 1985). By

analogy with the residence time of general particles with mass median diameters similar to that of

chromium, the residence time of atmospheric chromium is expected to be <10 days (Nriagu 1979). Based

on a troposphere to stratosphere turnover time of 30 years (EPA 1979), atmospheric particles with a

residence time of <10 days are not expected to transport from the troposphere to the stratosphere.

Since chromium compounds cannot volatilize from water, transport of chromium from water to the

atmosphere is not likely, except by transport in windblown sea sprays. Most of the chromium released

into water will ultimately be deposited in the sediment. A very small percentage of chromium can be

present in water in both soluble and insoluble forms. Soluble chromium generally accounts for a very

small percentage of the total chromium. Most of the soluble chromium is present as chromium(VI) and

soluble chromium(III) complexes. Less than 0.002% of total chromium in water and sediment in the

Amazon and Yukon Rivers was present in a soluble form (Cary 1982). In the aquatic phase,

chromium(III) occurs mostly as suspended solids adsorbed onto clayish materials, organics, or iron oxide

and Yukon rivers was in solution, the rest being present in the suspended solid phase (Cary 1982; King

1988). The ratio of suspended to dissolved solid in an organic-rich river in Brazil was 2.1 (Malm et al.

CHROMIUM 293

5. POTENTIAL FOR HUMAN EXPOSURE

1988). Soluble forms and suspended chromium can undergo intramedia transport. Chromium(VI) in

water will eventually be reduced to chromium(III) by organic matter in the water. It has been estimated

that the residence time of chromium (total) in Lake Michigan ranges from 4.6 to 18 years (Fishbein 1981;

Schmidt and Andren 1984).

(EPA 1980, 1984a; Fishbein 1981; Schmidt and Andren 1984). The bioavailability of chromium(III) to

1989). This decrease in bioavailability was attributed to lower availability of the free form of the metal

due to its complexation with humic acid. Based on this information, chromium is not expected to

biomagnify in the aquatic food chain. Although higher concentrations of chromium have been reported in

plants growing in high chromium-containing soils (e.g., soil near ore deposits or chromium-emitting

industries and soil fertilized by sewage sludge) compared with plants growing in normal soils, most of the

increased uptake in plants is retained in roots, and only a small fraction is translocated in the aboveground

part of edible plants (Cary 1982; WHO 1988). Therefore, bioaccumulation of chromium from soil

to above-ground parts of plants is unlikely (Petruzzelli et al. 1987). There is no indication of

biomagnification of chromium along the terrestrial food chain (soil-plant-animal) (Cary 1982).

in soil. A leachability study was conducted to study the mobility of chromium in soil. Due to different

pH values, a complicated adsorption process was observed and chromium moved only slightly in soil.

Chromium was not found in the leachate from soil, possibly because it formed complexes with organic

matter. These results support previous data finding that chromium is not very mobile in soil (Lin et al.

1996). These results are supported by leachability investigation in which chromium mobility was studied

for a period of 4 years in a sandy loam (Sheppard and Thibault 1991). The vertical migration pattern of

chromium in this soil indicated that after an initial period of mobility, chromium forms insoluble

complexes and little leaching is observed. Flooding of soils and the subsequent anaerobic decomposition

of plant detritus matters may increase the mobilization of chromium(III) in soils due to formation of

soluble complexes (Stackhouse and Benson 1989). This complexation may be facilitated by a lower soil

pH. A smaller percentage of total chromium in soil exists as soluble chromium(VI) and chromium(III),

which are more mobile in soil. The mobility of soluble chromium in soil will depend on the sorption

characteristics of the soil. The relative retention of metals by soil is in the order of lead > antimony >

CHROMIUM 294

5. POTENTIAL FOR HUMAN EXPOSURE

copper > chromium > zinc > nickel > cobalt > cadmium (King 1988). The sorption of chromium to soil

of soil. Chromium that is irreversibly sorbed onto soil, for example, in the interstitial lattice of geothite,

FeOOH, will not be bioavailable to plants and animals under any condition. Organic matter in soil is

1988). Chromium in soil may be transported to the atmosphere as an aerosol. Surface runoff from soil

can transport both soluble and bulk precipitate of chromium to surface water. Soluble and unadsorbed

chromium(VI) and chromium(III) complexes in soil may leach into groundwater. The leachability of

chromium(VI) in the soil increases as the pH of the soil increases. On the other hand, lower pH present in

acid rain may facilitate leaching of acid-soluble chromium(III) and chromium(VI) compounds in soil.

Chromium has a low mobility for translocation from roots to aboveground parts of plants (Cary 1982).

However, depending on the geographical areas where the plants are grown, the concentration of

chromium in aerial parts of certain plants may differ by a factor of 2–3 (Cary 1982).

In the atmosphere, chromium(VI) may be reduced to chromium(III) at a significant rate by vanadium

manganese oxide (EPA 1990b). However, this reaction is unlikely under most environmental conditions

(see Section 5.3.2.2). The estimated atmospheric half-life for chromium(VI) reduction to chromium(III)

was reported in the range of 16 hours to about 5 days (Kimbrough et al. 1999).

The reduction of chromium(VI) and the oxidation of chromium(III) in water has been investigated. The

half-life ranged from instantaneous to a few days. However, the reduction of chromium(VI) by organic

sediments and soils was much slower and depended on the type and amount of organic material and on

the redox condition of the water. The reaction was generally faster under anaerobic than aerobic

conditions. The reduction half-life of chromium(VI) in water with soil and sediment ranged from 4 to

140 days (Saleh et al. 1989). Dissolved oxygen by itself in natural waters did not cause any measurable

CHROMIUM 295

5. POTENTIAL FOR HUMAN EXPOSURE

oxidation of chromium(III) to chromium(VI) in 128 days (Saleh et al. 1989). When chromium(III) was

added to lake water, a slow oxidation of chromium(III) to chromium(VI) occurred, corresponding to an

oxidation half-life of nine years. Addition of 50 mg/L manganese oxide accelerated the process,

not be significant in most natural waters. The oxidation of chromium(III) to chromium(VI) during

chlorination of water was highest in the pH range of 5.5–6.0 (Saleh et al. 1989). However, the process

would rarely occur during chlorination of drinking water because of the low concentrations of

chromium(III) in these waters, and the presence of naturally occurring organics that may protect

chromium(III) from oxidation, either by forming strong complexes with chromium(III) or by acting as a

reducing agent to free available chlorine (EPA 1988c). In chromium(III)-contaminated waste waters

having pH ranges of 5–7, chlorination may convert chromium(III) to chromium(VI) in the absence of

chromium(III)-complexing and free chlorine reducing agents (EPA 1988c).

Chromium speciation in groundwater depends on the redox potential and pH conditions in the aquifer.

Chromium(VI) predominates under highly oxidizing conditions; whereas chromium(III) predominates

under reducing conditions. Oxidizing conditions are generally found in shallow aquifers, and reducing

conditions generally exist in deeper groundwaters. In sea water, chromium(VI) is generally stable (Fukai

oxidation state. This species and other chromium(III) species will predominate in more acidic pH;

The fate of chromium in soil is greatly dependent upon the speciation of chromium, which is a function of

redox potential and the pH of the soil. In most soils, chromium will be present predominantly in the

chromium(III) state. This form has very low solubility and low reactivity resulting in low mobility in the

environment and low toxicity in living organisms (Barnhart 1997). Under oxidizing conditions

relatively soluble, mobile, and toxic to living organisms. In deeper soil where anaerobic conditions exist,

chromium(VI) to chromium(III) is possible in aerobic soils that contain appropriate organic energy

sources to carry out the redox reaction. The reduction of chromium(VI) to chromium(III) is facilitated by

low pH (Cary 1982; EPA 1990b; Saleh et al. 1989). From thermodynamic considerations, chromium(VI)

CHROMIUM 296

5. POTENTIAL FOR HUMAN EXPOSURE

may exist in the aerobic zone of some natural soil. The oxidation of chromium(III) to chromium(VI) in

soil is facilitated by the presence of low oxidizable organic substances, oxygen, manganese dioxide, and

moisture. Oxidation is also enhanced at elevated temperatures in surface soil that result from brush fires

(Calder 1988; Cary 1982). Organic forms of chromium(III) (e.g., humic acid complexes) are more easily

oxidized than insoluble oxides. However, oxidation of chromium(III) to chromium(VI) was not observed

in soil under conditions of maximum aeration and a maximum pH of 7.3 (Bartlett and Kimble 1976). It

was later reported that soluble chromium(III) in soil can be partly oxidized to chromium(VI) by

manganese dioxide in soil, and the process is enhanced by pH higher than six (Bartlett 1991). Because

most chromium(III) in soil is immobilized due to adsorption and complexation with soil materials, the

barrier to this oxidation process is the lack of availability of mobile chromium(III) to immobile

manganese dioxide in soil surfaces. Due to this lack of availability of mobile chromium(III) to

manganese dioxide surfaces, a large portion of chromium in soil will not be oxidized to chromium(VI),

even in the presence of manganese dioxide and favorable pH conditions (Bartlett 1991; James et al.

1997).

The microbial reduction of chromium(VI) to chromium(III) has been discussed as a possible remediation

technique in heavily contaminated environmental media or wastes (Chen and Hao 1998). Factors

affecting the microbial reduction of chromium(VI) to chromium(III) include biomass concentration,

initial chromium(VI) concentration, temperature, pH, carbon source, oxidation-reduction potential and the

presence of both oxyanions and metal cations. Although high levels of chromium(VI) are toxic to most

microbes, several resistant bacterial species have been identified which could ultimately be employed in

remediation strategies (Chen and Hao 1998). Elemental iron, sodium sulfite, sodium hydrosulfite, sodium

bisulfite, sodium metabisulfite sulfur dioxide and certain organic compounds such as hydroquinone have

also been shown to reduce chromium(VI) to chromium(III) and have been discussed as possible

remediation techniques in heavily contaminated soils (James et al. 1997; Higgins et al. 1997). The

limitations and efficacy of these and all remediation techniques are dependent upon the ease in which the

reducing agents are incorporated into the contaminated soils.

Reliable evaluation of the potential for human exposure to chromium depends in part on the reliability of

supporting analytical data from environmental samples and biological specimens. In reviewing data on

chromium levels monitored or estimated in the environment, it should also be noted that the amount of

chemical identified analytically is not necessarily equivalent to the amount that is bioavailable.

CHROMIUM 297

5. POTENTIAL FOR HUMAN EXPOSURE

urban and nonurban areas during 1977–1984 are reported in EPA (1990b). Levels of total chromium in

the ambient air in U.S. urban and nonurban areas during 1977–1984 are reported in EPA's National

Aerometric Data Bank (EPA 1984a, 1990b). According to this databank, the arithmetic mean total

locations that showed the highest total arithmetic mean chromium concentrations were in Steubenville,

1990b). An indoor/outdoor air study was conducted in southwestern Ontario to measure levels of

chromium(VI) and the size fraction of chromium(VI). Indoor and outdoor samples were taken from

concentrations were less than ½ of the outdoor concentrations. Analysis of airborne chromium(VI)

particles showed that they are inhalable in size (Bell and Hipfner 1997). During the period 1978–1982,

The maximum level of total chromium in ambient air samples in Corpus Christi, Texas, a site of chromate

recent study measured the levels of chromium(VI) and total chromium in the ambient air. The

concentrations of chromium(VI) in the indoor air of residences in Hudson County, New Jersey, in 1990

study was conducted at 25 industrial sites in Hudson County, New Jersey to analyze soils containing

chromite ore processing residues. The industrial sites include industrial, manufacturing, trucking, and

warehouse facilities. The study found industrial indoor chromium(VI) and total chromium concentrations

levels of chromium(VI) in soil do not result in higher levels of chromium(VI) in air (Finley et al. 1993).

CHROMIUM 298

5. POTENTIAL FOR HUMAN EXPOSURE

(Scott et al. 1997a). The mean airborne chromium(VI) and total chromium concentrations in the indoor

air of industrial sites in Hudson County, New Jersey, contaminated by chromite ore-processing residue

airborne chromium(VI) and total chromium concentrations in outdoor air for the same sites were

1992). An air dispersion model was recently developed which accurately estimated chromium(VI)

concentrations at two of these industrial sites in Hudson County, NJ (Scott et al. 1997b). The background

0.0 (0.41); 6.2 (7.7); 0.9 (1.7); 2.8 (2.7); 0.0 (0.08); 0.3 (0.1); and 1.2 (0.12). The estimated percent levels

of chromium(III) and chromium(VI) in the U.S. atmosphere from anthropogenic sources are given in

Table 5-1 (EPA 1990b). Fly ash from a coal-fired power plant contained 1.4–6.1 mg/kg chromium(VI)

(Stern et al. 1984). In a field study to assess inhalation exposure to chromium during showering and

bathing activities, the average chromium(VI) concentration in airborne aerosols ranged from 87 to

shower (Finley et al. 1996a).

Hoff 1985; Cary 1982; Schroeder et al. 1987; Sheridan and Zoller 1989). Saltzman et al. (1985)

compared the levels of atmospheric chromium at 59 sites in U.S. cities during 1968–1971 with data from

EPA's National Aerometric Data Bank file for 1975–1983. They concluded that atmospheric chromium

levels may have declined in the early 1980s from the levels detected in the 1960s and 1970s.

Chromium concentrations in U.S. river water usually range from <1 to 30 µg/L (EPA 1984a; Malm et al.

1988; Ramelow et al. 1987), with a median value of 10 µg/L (Eckel and Jacob 1988; Smith et al. 1987).

Chromium concentrations in lake water generally do not exceed 5 µg/L (Borg 1987; Cary 1982). The

higher levels of chromium can be related to source(s) of anthropogenic pollution. Dissolved chromium

concentrations of 0.57–1.30 µg/L were reported in the Delaware River near Marcus Hook and Fieldsboro,

Pennsylvania in January 1992, with chromium(III) composing 67% of the total (Riedel and Sanders

1998). In March 1992 these concentrations decreased to 0.03–0.23 µg/L. The chromium levels detected

in drinking water in an earlier study (1962–1967 survey) may be erroneous due to questionable sampling

and analytical methods (see Section 6.1) (EPA 1984a). A survey conducted from 1974 to 1975 that had a

detection limit of 0.1 µg/L and surveyed 3,834 U.S. tap waters probably provides better estimates of

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chromium concentrations in U.S. drinking water. The survey reported chromium concentrations in

drinking water that ranged from 0.4 to 8.0 µg/L, with a mean value of 1.8 µg/L (Greathouse and Craun

1978). These values may be higher than the actual values, due to inadequate flushing of tap water before

sample collection (EPA 1984a). The concentration of chromium in household tap water may be higher

than supply water due to corrosion of chromium-containing pipes. At a point of maximum contribution

from corrosion of the plumbing system, the peak chromium in tap water in Boston, Massachusetts, was

15 µg/L (Ohanian 1986). A survey that targeted drinking waters from 115 Canadian municipalities

during 1976–1977 reported the median and the range of chromium concentrations to be <2.0 µg/L

(detection limit) and <2.0–8.0 µg/L, respectively (Meranger et al. 1979). The mean chromium

concentration in ocean water is 0.3 µg/L, with a range of 0.2–50 µg/L (Cary 1982). In general, the

concentration of chromium in ocean water is much lower than that in lakes and rivers. The concentrations

of total chromium in groundwater at the Idaho National Engineering Laboratory, where chromate is used

as a corrosion inhibitor, ranged from <1 to 280 µg/L (USGS 1989). The water from a village well

situated near a waste pond receiving chromate waste in Douglas, Michigan, contained 10,800 µg/L

chromium(VI). Similarly, water from a private well adjacent to an aircraft plant in Nassau County, New

York, contained 25,000 µg/L chromium(VI), while water from a public well adjacent to another aircraft

plant in Bethpage, New York, contained 1,400 µg/L (chromium(VI) (Davids and Lieber 1951). In a later

study, water from an uncontaminated well in Nassau County, New York, contained an undetectable level

of chromium(VI), whereas a contaminated well in the vicinity of a plating plant contained 6,000 µg/L

chromium(VI) (Lieber et al. 1964). A high chromium concentration (120 µg/L) was detected in private

drinking water wells adjacent to a NPL site in Galena, Kansas (ATSDR 1990a). The mean concentration

of chromium in rainwater is 0.14–0.9 µg/L (Barrie et al. 1987; Dasch and Wolff 1989).

The chromium level in soils varies greatly and depends on the composition of the parent rock from which

the soils were formed. Basalt and serpentine soils, ultramafic rocks, and phosphorites may contain

chromium as high as a few thousand mg/kg (Merian 1984) whereas soils derived from granite or

sandstone will have lower concentrations of chromium (Swaine and Mitchell 1960). The concentration

range of chromium in 1,319 samples of soils and other surficial materials collected in the conterminous

United States was 1–2,000 mg/kg, with a geometric mean of 37 mg/kg (USGS 1984). Chromium

concentrations in Canadian soils ranged from 5 to 1,500 mg/kg, with a mean of 43 mg/kg (Cary 1982). In

a study with different kinds of soils from 20 diverse sites including old chromite mining sites in

Maryland, Pennsylvania, and Virginia, the chromium concentration ranged from 4.9 to 71 mg/kg (Beyer

CHROMIUM 300

5. POTENTIAL FOR HUMAN EXPOSURE

and Cromartie 1987). A polynuclear aromatic hydrocarbon (PAH) soil study was conducted to determine

the metal levels in soil at the edge of a busy road that runs through the Aplerbecker Forest in West

Germany. Chromium(VI) concentrations of 64 mg/kg were measured, and these concentrations were 2-

to 4-fold higher along the road than in the natural forest (Munch 1993). The soil beneath decks treated

with chrominated copper arsenate (CCA), a wood preservative, had an average chromium content of

43 mg/kg (Stilwell and Gorny 1997).

Chromium has been detected at a high concentration (43,000 mg/kg) in soil at the Butterworth Landfill

site in Grand Rapid City, Michigan, which was a site listed on the NPL (ATSDR 1990b).

Chromium was detected in sediment obtained from the coastal waters of the eastern U.S. seashore at

concentrations of 3.8–130.9 µg/g in 1994 and 0.8–98.1 µg/g in 1995 (Hyland et al. 1998).

The concentration of chromium in the particulate portion of melted snow collected from two urban areas

(Toronto and Montreal) of Canada ranged from 100 to 3,500 mg/kg (Landsberger et al. 1983). In the

suspended materials and sediment of water bodies, chromium levels ranged from 1 to 500 mg/kg (Byrne

and DeLeon 1986; EPA 1984a; Mudroch et al. 1988; Ramelow et al. 1987). The chromium concentration

in incinerated sewage sludge ash may be as high as 5,280 mg/kg (EPA 1984a).

Total chromium levels in most fresh foods are extremely low (vegetables (20–50 µg/kg), fruits

(20 µg/kg), and grains and cereals (40 µg/kg)) (Fishbein 1984). The chromium levels of various foods

are reported in Table 5-3. In a study to find the concentrations of chromium in edible vegetables in

Tarragon Province, Spain, the highest levels of chromium were found in radish root and spinach, with a

non-significant difference between the samples collected in two areas (northern industrial and southern

agricultural). The samples ranged in concentration from 0.01 µg/g to 0.21 µg/g (industrial) and from

0.01 µg/g to 0.22 µg/g (agricultural) (Schuhmaker et al. 1993). Acidic foods that come into contact with

stainless steel surfaces during harvesting, processing, or preparation for market are sometimes higher in

chromium content because of leaching conditions. Processing, however, removes a large percentage of

chromium from foods (e.g., whole-grain bread contains 1,750 µg/kg chromium, but processed white

bread contains only 140 µg/kg; and molasses contains 260 µg/kg chromium, but refined sugar contains

only 20 µg/kg chromium) (Anderson 1981; EPA 1984a).

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Sample

Mean concentration

(µg/kg) Reference

Fresh vegetables 30–140 EPA 1984a

Frozen vegetables 230 EPA 1984a

Canned vegetables 230 EPA 1984a

Fresh fruits 90–190 EPA 1984a

Fruits 20 EPA 1984a

Canned fruits 510 EPA 1984a

Dairy products 100 EPA 1984a

Chicken eggs 160–520 Kirpatrick and Coffin 1975

Chicken eggs 60 Kirpatrick and Coffin 1975

Whole fish 50–80 EPA 1984a

Edible portion of fresh fin fish <100–160 Eisenberg and Topping 1986

Meat and fish 110–230 EPA 1984a

Seafoods 120–470 EPA 1984a

Grains and cereals 40–220 EPA 1984a

CHROMIUM 302

5. POTENTIAL FOR HUMAN EXPOSURE

Chromium levels in oysters, mussels, clams, and mollusks vary from <0.1 to 6.8 mg/kg (dry weight)

(Byrne and DeLeon 1986; Ramelow et al. 1989). Fish and shellfish collected from ocean dump sites off

New York City, Delaware Bay, and New Haven, Connecticut, contained <0.3–2.7 mg/kg chromium (wet

weight) (Greig and Jones 1976). The chromium concentration in fish sampled from 167 lakes in the

northeastern United States was 0.03–1.46 µg/g with a mean concentration of 0.19 µg/g (Yeardley et al.

1998). Higher levels of chromium in forage of meat animals have been reported for plants grown in soils

with a high concentration of chromium (see Section 5.3.1). Cigarette tobacco reportedly contains

0.24–14.6 mg/kg chromium, but no estimates were available regarding the chromium levels in inhaled

cigarette smoke (Langård and Norseth 1986). Cigarette tobacco grown in the United States contains

Cement-producing plants are a potential source of atmospheric chromium. Portland cement contains

41.2 mg/kg chromium (range 27.5–60 mg/kg). Soluble chromium accounts for 4.1 mg/kg (range

1.6–8.8 mg/kg) of which 2.9 mg/kg (range 0.03–7.8 mg/kg) is chromium(VI) (Fishbein 1981). The

wearing down of vehicular brake linings that contain asbestos represents another source of atmospheric

on U.S. automobiles in 1975 in the United States represented an additional source of atmospheric

centimeter in vehicular exhaust, under various operating conditions (Fishbein 1981).

The general population is exposed to chromium by inhaling ambient air, ingesting food, and drinking

water containing chromium. Dermal exposure of the general public to chromium can occur from skin

contact with certain consumer products that contain chromium. Some of the consumer products known to

contain chromium are certain wood preservatives, cement, cleaning materials, textiles, and leather tanned

with chromium (WHO 1988). However, no quantitative data for dermal exposure to chromiumcontaining

(Fishbein 1984) and tap water (<2 µg/L) (Greathouse and Craun 1979) have been used to estimate the

daily intake of chromium via inhalation (<0.2–0.6 µg) and via tap water (<4 µg). These estimates are

chromium intake for the U.S. population from consumption of selected diets (diets with 25 and 43% fat)

has been estimated to range from 25 to 224 µg with an average of 76 µg (Kumpulainen et al. 1979). The

average value is close to a value of 60 µg reported by Bennett (1986). The bioavailability of chromium

CHROMIUM 303

5. POTENTIAL FOR HUMAN EXPOSURE

from different foods may vary. No correlation was found between the insulin potentiation and the total

chromium extractable from foods by acid hydrolysis. However, a significant correlation was found

between the ethanol-extractable chromium and biological activity. The highest amounts of ethanolextractable

chromium were found in brewer's yeast, black pepper, calf-liver, cheese, and wheatgerm

(WHO 1988). The chromium concentrations in tissues and body fluids of the general population are

given in Table 5-4.

Workers in industries that use chromium can be exposed to concentrations of chromium two orders of

magnitude higher than exposure to the general population (Hemminki and Vainio 1984). Occupational

exposure to chromium occurs mainly from chromate production, stainless steel production and welding,

chromium plating, ferrochrome alloys and chrome pigment production, and working in tanning industries.

A list of industries that may be sources of chromium exposure is given in Table 5-5. For most occupations,

exposure is due to both chromium(III) and chromium(VI) present as soluble and insoluble fractions.

However, exceptions include: the tanning industry, where exposure is mostly from soluble chromium(III)

and the plating industry, where exposure is due to soluble chromium(VI). The typical concentration ranges

of airborne chromium(VI) to which workers in these industries were exposed during an average of

occupational airborne chromium concentrations have declined significantly since the 1970s (Stern 1982).

In a recent study conducted in Taiwan to estimate worker exposure to chromium in electroplating factories,

manufacturing area (Kuo et al. 1997b). In an occupational exposure study of chromium in an aircraft

construction factory, airborne samples were collected over a 4-hour period; urinary samples were collected

at the beginning (Monday), end (Friday), and after the work shift in order to analyze the absorption of

the urine sampling results were 0.16–7.74 µg/g creatinine. Compared to the ACGIH and BEI-ACGIH

Occupational Exposure Survey (NOES) conducted by NIOSH from 1981 to 1983 estimated that

304,829 workers in the United States were potentially exposed to chromium(VI) (NIOSH 1989). The

NOES database does not contain information on the frequency, concentration, or

CHROMIUM 304

5. POTENTIAL FOR HUMAN EXPOSURE

Sample Median/mean Range Reference

Serum 0.006 µg/L 0.01–0.17 µg/L Sunderman et al. 1987

Urine 0.4 µg/L 0.24–1.8 µg/L Iyengar and Woittiez 1988

Lung 201 µg/kg (wet weight) 28–898 µg/kg (wet weight) Raithel et al. 1987

Breast milk 0.30 µg/L 0.06–1.56 µg/L Casey and Hambidge 1984

Hair 0.234 mg/kg Not available Takagi et al. 1986

Nail 0.52 mg/kg No applicable Takagi et al. 1988

CHROMIUM 305

5. POTENTIAL FOR HUMAN EXPOSURE

Abrasives manufacturers

Acetylene purifiers

Adhesives workers

Aircraft sprayers

Alizarin manufacturers

Alloy manufactures

Aluminum anodizers

Anodizers

Battery manufacturers

Biologists

Blueprint manufacturers

Boiler scalers

Candle manufacturers

Cement workers

Ceramic workers

Chemical workers

Chromate workers

Chromium-alloy workers

Chromium-alum workers

Chromium platers

Copper etchers

Copper-plate strippers

Corrosion-inhibitor workers

Crayon manufacturers

Diesel locomotive repairmen

Drug manufacturers

Dye manufacturers

Dyers

Electroplaters

Enamel workers

Explosive manufacturers

Fat purifiers

Fireworks manufacturers

Flypaper manufacturers

Furniture polishers

Fur processors

Glass-fibre manufacturers

Glue manufacturers

Histology technicians

Jewelers

Laboratory workers

Leather finishers

Linoleum workers

Lithographers

Magnesium treaters

Match manufacturers

Metal cleaners

Metal workers

Milk preservers

Oil drillers

Oil purifiers

Painters

Palm-oil bleachers

Paper water proofers

Pencil manufacturers

Perfume manufacturers

Photoengravers

Photographers

Platinum polishers

Porcelain decorators

Pottery frosters

Pottery glazers

Printers

Railroad engineers

Refractory-brick manufacturers

Rubber manufacturers

Shingle manufacturers

Silk-screen manufacturers

Smokeless-powder manufacturers

Soap manufacturers

Sponge bleachers

Steel workers

Tanners

Textile workers

Wallpaper printers

Wax workers

Welders

Wood-preservative workers

Wood stainers

CHROMIUM 306

5. POTENTIAL FOR HUMAN EXPOSURE

duration of exposure; the survey only estimates the number of workers potentially exposed to chemicals

in the workplace.

In a survey of workers in pigment factories in England that produced strontium and lead chromate, the

concentrations of chromium in the whole blood in exposed workers ranged from 3 to 216 µg/L, compared

to a level of <1 µg/L for the nonoccupationally exposed population (McAughey et al. 1988). The

corresponding concentrations in the urine of exposed workers and the unexposed population were

1.8–575 µg chromium/g creatinine and <0.5 µg chromium/g creatinine, respectively (McAughey et al.

1988). Other investigators have found a higher lung burden for chromium in occupational groups than in

unexposed groups. The median concentration of chromium in the lungs of deceased smelter workers in

Sweden was 450 µg/kg (wet weight), compared to a value of 110 µg/kg (wet weight) for rural controls

and 199 µg/kg (wet weight) for urban controls (Gerhardsson et al. 1988).

This section focuses on exposures from conception to maturity at 18 years in humans. Differences from

adults in susceptibility to hazardous substances are discussed in Section 2.7, Children’s Susceptibility.

Children are not small adults. A child’s exposure may differ from an adult’s exposure in many ways.

Children drink more fluids, eat more food, breathe more air per kilogram of body weight, and have a

larger skin surface in proportion to their body volume. A child’s diet often differs from that of adults.

The developing human’s source of nutrition changes with age: from placental nourishment to breast milk

or formula to the diet of older children who eat more of certain types of foods than adults. A child’s

behavior and lifestyle also influence exposure. Children crawl on the floor, put things in their mouths,

sometimes eat inappropriate things (such as dirt or paint chips), and spend more time outdoors. Children

also are closer to the ground, and they do not use the judgment of adults to avoid hazards (NRC 1993).

Children living in vicinities where there are chromium waste sites nearby may be exposed to chromium to

a greater extent than adults through inhalation of chromium particulates and through contact with

contaminated soils. One study has shown that the average concentration of chromium in the urine of

children at ages five and younger was significantly higher than in adults residing near sites where

chromium waste slag was used as fill material (Fagliano et al. 1997). The tendency of young children to

ingest soil, either intentionally through pica or unintentionally through hand-to-mouth activity, is well

documented. These behavioral traits can result in ingestion of chromium present in soil and dust. Soil

CHROMIUM 307

5. POTENTIAL FOR HUMAN EXPOSURE

may affect the bioavailability of contaminants in several ways. The most likely manner is by acting as a

competitive sink for the contaminants. In the presence of soil, the contaminants will partition between

absorption by the gut and sorption onto the soil particles. If a soil has a longer residence time in the gut

than food particles, sorption may enhance the overall absorption of the contaminant (Sheppard et al.

1995). If the contaminant is irreversibly bound to soil particles, the contaminant is unlikely to be

absorbed in the gastrointestinal tract. Hexavalent chromium exists in soils as a relatively soluble anion

and may be present in bioavailable form. In contrast, chromium(III) present in soil is generally not very

soluble or mobile under most environmental conditions and is not readily bioavailable (James et al. 1997).

Studies discussing the oral absorption of chromium in rats from a soil surface in which 30% of the

chromium was in hexavalent form and 70% was in trivalent form suggested that while absorption in

animals is quite low, chromium appeared to be better absorbed from soil than from soluble chromate salts

(Witmer et al. 1989, 1991). However, less than half of the administered dose of chromium could be

accounted for in this study, and in separate experiments with low dosages administered to the rats, the

control animals actually had higher concentrations of chromium than the animals that were administered

the oral dose. Children may accidently ingest chromium picolinate in households whose members use

this product as a dietary supplement unless it is well stored and kept away from children. Small amounts

of chromium are used in certain consumer products such as toners in copying machines and printers, but

childhood exposure from these sources are expected to be low. Children may also be exposed to

chromium from parents’ clothing or items removed from the workplace if the parents are employed in a

setting where occupational exposure is significant (see Section 5.5). Chromium has been detected in

breast milk at concentrations of 0.06–1.56 µg/L (Casey and Hambidge 1984), suggesting that children

could be exposed to chromium from breast-feeding mothers. Studies on mice have shown that chromium

crosses the placenta and can concentrate in fetal tissue (Danielsson et al. 1982; Saxena et al.1990a).

In addition to individuals who are occupationally exposed to chromium (see Section 5.5), there are

several groups within the general population that have potentially high exposures (higher than

background levels) to chromium. These populations include individuals living in proximity to sites

where chromium was produced or sites where chromium was disposed, and individuals living near one

of the NPL hazardous waste sites where chromium has been detected in some environmental media

(HazDat 2000). Persons using chromium picolinate as a dietary supplement will also be exposed to

higher levels of chromium than those not ingesting this product (Anderson 1998b). Like many other

products used to promote weight loss or speed metabolism there is also the potential for overuse of this

CHROMIUM 308

5. POTENTIAL FOR HUMAN EXPOSURE

product by some members of the population in order to achieve more dramatic results (Wasser et al.

1997). People may also be exposed to higher levels of chromium if they use tobacco products, since

tobacco contains chromium (IARC 1980).

Workers in industries that use chromium are one segment of the population that is especially at high risk

to chromium exposure. Occupational exposure from chromate production, stainless steel welding,

chromium plating, and ferrochrome and chrome pigment production is especially significant since the

exposure from these industries is to chromium(VI). Occupational exposure to chromium(III) compounds

may not be as great a concern as exposure to chromium(VI) compounds. Among the general population,

residents living near chromate production sites may be exposed to higher levels of chromium(VI) in air.

Baltimore, Maryland (EPA 1984a). People who live near chromium waste disposal sites and chromium

manufacturing and processing plants may be exposed to elevated levels of chromium. The airborne

concentrations of chromium(VI) and total chromium in a contaminated site in Hudson County, New

Jersey, were studied (Falerios et al. 1992). The mean concentrations of both chromium(VI) and total

chromium in indoor air of the contaminated site were about three times higher than the mean indoor air

concentrations of uncontaminated residential sites in Hudson County. Although the mean concentration

of chromium(VI) in outdoor air was much lower than the current occupational exposure limit of

for urban New Jersey. However, recent sampling data from Hudson County, NJ have shown that more

than two-thirds of previously sampled sites contaminated with chromite ore processing residue did not

have statistically significant mean concentrations greater than the background levels (Scott et al. 1997a).

This data, as well as the results of a soil dispersion model (Scott et al. 1997b) suggest that heavy

vehicular traffic over unpaved soil surfaces containing chromium(VI) are required for high levels of

atmospheric chromium(VI) at these sites. Persons using contaminated water for showering and bathing

activities may also be exposed via inhalation to potentially high levels of chromium(VI) in airborne

aerosols (Finley et al. 1996a). In a field study to simulate daily bathing activity, airborne chromium(VI)

concentrations were about 2 orders of magnitude greater than ambient outdoor air concentrations when

water concentrations of 5.4 and 11.5 mg/L were used in the shower.

A study was conducted from September to November 1989 to determine the levels of chromium in urine

and red blood cells of state employees who worked at a park (with only indirect exposure potential)

CHROMIUM 309

5. POTENTIAL FOR HUMAN EXPOSURE

adjacent to chromium-contaminated sites in Hudson County, New Jersey (Bukowski et al. 1991). The

chromium levels in red blood cells and urine of 17 of these employees showed no differences compared

to 36 employees who worked at state parks outside Hudson County. The authors concluded that urinary

and blood levels of chromium are poor biological markers in gauging low-level environmental exposure

to chromium. This study also concluded that chromium levels in blood and urine depended on other

confounding variables, such as exercise, past employment in a chromium-exposed occupation, beer

drinking, and diabetic status. Other lifestyle (e.g., smoking), dietary, or demographic factors had no

measurable effect on blood and urinary chromium. These conclusions are consistent with the results of a

recent study that measured the urinary excretion of chromium following oral ingestion of chromite ore

processing residue material for three days (Finley and Paustenbach 1997). These results indicate no

statistical difference in mean urinary chromium concentrations in groups of individuals exposed to

chromite ore processing residue material versus the control group. High levels of chromium were

detected in the urine and hair of individuals living near a chromite ore-processing plant in Mexico

(Rosas et al. 1989), and does suggest the possibility of using these media as biological markers in

gauging long term high-level environmental exposure to chromium.

Elevated levels of chromium in blood, serum, urine, and other tissues and organs have been observed in

patients with cobalt-chromium knee and hip arthroplasts (Coleman et al. 1973; Michel et al. 1987;

Sunderman et al. 1989).

The chromium content in cigarette tobacco from the United States has been reported to be

0.24–6.3 mg/kg (IARC 1980), but neither the chemical form nor the amount of chromium in tobacco

smoke is known. People who use tobacco products may be exposed to higher than normal levels of

chromium.

Section 104(i)(5) of CERCLA, as amended, directs the Administrator of ATSDR (in consultation with

the Administrator of EPA and agencies and programs of the Public Health Service) to assess whether

adequate information on the health effects of chromium is available. Where adequate information is not

available, ATSDR, in conjunction with the National Toxicology Program (NTP), is required to assure

the initiation of a program of research designed to determine the health effects (and techniques for

developing methods to determine such health effects) of chromium.

CHROMIUM 310

5. POTENTIAL FOR HUMAN EXPOSURE

The following categories of possible data needs have been identified by a joint team of scientists from

ATSDR, NTP, and EPA. They are defined as substance-specific informational needs that if met would

reduce the uncertainties of human health assessment. This definition should not be interpreted to mean

that all data needs discussed in this section must be filled. In the future, the identified data needs will be

evaluated and prioritized, and a substance-specific research agenda will be proposed.

properties of chromium and its compounds are known (Hartford 1979; Weast 1985) and prediction of

environmental fate and transport of chromium in environmental media is possible. However, the physical

or chemical forms and the mode by which chromium(III) compounds are incorporated into biological

systems are not well characterized. The determination of the solubilities of hexavalent chromium

compounds in relevant body fluids (e.g., the solubility of chromates in lung fluid) may also be helpful.

production volume is important because it may indicate environmental contamination and human

exposure. If a chemical's production volume is high, there is an increased probability of general

population exposure via consumer products and environmental sources, such as air, drinking water, and

food. Data concerning the production (Hartford 1979; SRI 1997), import (NTDB 1998), and use (CMR

1988; EPA 1984a; IARC 1990; USDI 1988a) of commercially significant chromium compounds are

available. It does not appear that chromium is used to process foods for human consumption or is added

to foods other than diet supplements. Consumer exposure to chromium occurs mostly from natural food

(Bennett 1986; EPA 1984a; Kumpulainen et al. 1979), but this exposure will increase particularly for

people who consume acidic food cooked in stainless steel utensils (Anderson 1981; EPA 1984a).

Exposure to chromium occurs to a much lesser extent from products such as toners of photocopying

machines, some wood treatment chemicals, and through other chromium-containing consumer products

(CMR 1988; EPA 1984a; IARC 1990; USDI 1988a).

The TRI97 (1999), which became available in 1999, has been used in this profile. As can be seen in

Table 5-2, the largest amount of chromium from production and user facilities is disposed of on land or

transferred to an off-site location. More detailed site-and medium-specific (e.g., air, water, or soil) release

data for chromium that is disposed of off-site are necessary to assess the exposure potential to these

compounds from different environmental media and sources. There are EPA guidelines regarding the

CHROMIUM 311

5. POTENTIAL FOR HUMAN EXPOSURE

disposal of chromium wastes and OSHA regulations regarding the levels of chromium in workplaces

(EPA 1988a; OSHA 1998a).

According to the Emergency Planning and Community Right-to-Know Act of 1986, 42 U.S.C.

Section 11023, industries are required to submit chemical release and off-site transfer information to the

EPA. The Toxics Release Inventory (TRI), which contains this information for 1997, became available in

1999. This database will be updated yearly and should provide a list of industrial production facilities

and emissions.

transport of chromium in air (Schroeder et al. 1987; Scott et al. 1997a, 1997b), water (Carey 1982; EPA

1980, 1984a; Fishbein 1981; Schmidt and Andren 1984) and soil (Bartlett 1991; Calder 1988; Cary

1982). Chromium is primarily removed from the atmosphere by fallout and precipitation. By analogy

with copper, the residence time of chromium in the atmosphere is expected to be <10 days (Nriagu 1979).

Most of the chromium in lakes and rivers will ultimately be deposited in the sediments. Chromium in the

aquatic phase occurs in the soluble state or as suspended solids adsorbed onto clayish materials, organics,

or iron oxides (Cary 1982). Most of the soluble chromium is present as chromium(VI) or as soluble

chromium(III) complexes and generally accounts for a small percentage of the total (Cary 1982).

Additional data, particularly regarding chromium's nature of speciation, would be necessary to fully

assess chromium's fate in air. For example, if chromium(III) oxide forms some soluble salt in the air due

to speciation, its removal by wet deposition will be faster. No data regarding the half-life of chromium in

the atmosphere or a measure of its persistence are available. In aquatic media, sediment will be the

ultimate sink for chromium, although soluble chromates may persist in water for years (Cary 1982; EPA

1984a). Additional data elucidating the nature of speciation of chromium in water and soil would also be

desirable.

contaminated air, water, soil, or plant material in the environment has not been adequately studied.

Absorption studies of chromium in humans and animals provide information regarding the extent and rate

of inhalation (Cavalleri and Minoia 1985; Kiilunen et al. 1983; Langård et al. 1978), and oral exposure

(Anderson 1981, 1986; Anderson et al. 1983; Donaldson and Barreras 1966; Randall and Gibson 1987;

Suzuki et al. 1984). These studies indicate that chromium(VI) compounds are generally more readily

absorbed from all routes of exposure than are chromium(III) compounds. This is consistent, in part, with

the water solubilities of these compounds (Bragt and van Dura 1983). The bioavailability of both forms

CHROMIUM 312

5. POTENTIAL FOR HUMAN EXPOSURE

is greater from inhalation exposure than from ingestion or dermal exposure. The bioavailability of

chromium from soil depends upon several factors (Witmer et al. 1989). Factors that may increase the

mobility of chromium in soils include the speculated conversion of chromium(III) to chromium(VI),

increases in pH, and the complexation of chromium(III) with organic matter from water-soluble

complexes. Data on the bioavailability of chromium compounds from actual environmental media and

the difference in bioavailability for different media need further development.

Fishbein 1981; Schmidt and Andren 1984). There is no indication of biomagnification of chromium

along the aquatic food chain (Cary 1982). Some data indicate that chromium has a low mobility for

translocation from roots to above-ground parts of plants (Cary 1982; WHO 1988). However, more data

regarding the transfer ratio of chromium from soil to plants and biomagnification in terrestrial food chains

would be desirable.

chromium concentrations in U.S. drinking water typically range from 0.4 to 8.0 µg/L, with a mean value

of 1.8 µg/L (Greathouse and Craun 1978). The chromium level in soils varies greatly and depends on the

composition of the parent rock from which the soils were formed. Basalt and serpentine soils, ultramafic

rocks, and phosphorites may contain chromium as high as a few thousand mg/kg (Merian 1984), whereas

soils derived from granite or sandstone will have lower concentrations of chromium (Swaine and Mitchell

1960). The concentration range of chromium in 1,319 samples of soils and other surficial materials

collected in the conterminous United States was 1–2,000 mg/kg, with a geometric mean of 37 mg/kg

(USGS 1984). There is a large variation in the available data regarding the levels of chromium in foods

(EPA 1984a). Concentrations ranges are 30–230 µg/kg in vegetables, 20–510 µg/kg in fruits,

40–220 µg/kg in grains and cereals, and 110–230 µg/kg in meats and fish (EPA 1984a). It would be

useful to develop nationwide monitoring data on the levels of chromium in U.S. ambient air and drinking

water, and these data should quantitate levels of both chromium(III) and chromium(VI) and not just total

chromium.

Reliable monitoring data for the levels of chromium in contaminated media at hazardous waste sites are

needed so that the information obtained on levels of chromium in the environment can be used in

combination with the known body burdens of chromium to assess the potential risk of adverse health

effects in populations living in the vicinity of hazardous waste sites.

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air and ingesting food and drinking water containing chromium. Dermal exposure of the general public to

chromium can occur from skin contact with certain consumer products that contain chromium or from

contact with chromium contaminated soils. Some of the consumer products known to contain chromium

are certain wood preservatives, cement, cleaning materials, textiles, and leather tanned with chromium

(WHO 1988). However, no quantitative data for dermal exposure to chromium-containing consumer

water (<2 µg/L) (Greathouse and Craun 1979) have been used to estimate the daily intake of chromium

via inhalation (<0.2–0.6 µg) and via tap water (<4 µg). These estimates are based on an air inhalation rate

population from consumption of selected diets (diets with 25 and 43% fat) has been estimated to range

from 25 to 224 µg, with an average of 76 µg (Kumpulainen et al. 1979). The average value is close to a

value of 60 µg reported by Bennett (1986). However, few data on the levels of chromium in body tissues

or fluids for populations living near hazardous waste sites are available. Such data could be a useful tool

as an early warning system against harmful exposures. In addition, there is a need for data on the

background levels of chromium in body fluids of children. Such data would be important in assessing the

exposure levels of this group of people.

This information is necessary for assessing the need to conduct health studies on these populations.

children. Chromium has been detected in breast milk at concentrations of 0.06–1.56 µg/L (Casey and

Hambidge 1984), suggesting that children could be exposed to chromium from breast-feeding mothers.

Studies in mice have shown that chromium crosses the placenta and can concentrate in fetal tissue

(Danielsson et al. 1982; Saxena et al.1990a). Because children living near areas contaminated with

chromium have been shown to have elevated chromium levels in urine as compared to adults (Fagliano et

al. 1997), additional body burden studies are required to evaluate the exposures and the potential

consequences this might have upon children. This is particularly important around heavily contaminated

soils where children may be exposed dermally or through inhalation of soil particulates during play

activities. These studies may determine if children may be more susceptible than adults to the toxic

effects of chromium including immunosensitivity. Studies are necessary that examine children’s weightadjusted

intake of chromium and determine how it compares to that of adults. Since chromium is often

detected in soil surfaces and children ingest soil either intentionally through pica or unintentionally

through hand-to-mouth activity, pica is a unique exposure pathway for children. Studies have shown that

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although absorption of chromium is low, it may be enhanced slightly from contaminated soil surfaces

(Witmer et al. 1989, 1991).

Child health data needs relating to susceptibility are discussed in 2.12.2 Identification of Data Needs:

Children’s Susceptibility.

established in the National Exposure Registry. The information that is amassed in the National Exposure

Registry facilitates the epidemiological research needed to assess adverse health outcomes that may be

related to the exposure to chromium.

The database of federal research programs in progress (FEDRIP) indicates several current projects that

may fill some existing data gaps. Dr. Syed M. Naqvi at Southern University is investigating the extent to

which chromium can bioaccumulate in crayfish tissues consumed by humans (FEDRIP 1999).

Dr. Gordon E. Brown, Jr. at Stanford University is examining the possibility of using the results of

spectroscopic measurements in developing a predictive macroscopic model of sorption behavior and

fluid-solid partitioning of chromium (FEDRIP 1999). Dr. Kathleen Dixon at the University of Cincinnati

is investigating the role of the biotransformation of chromium within the cell in the mutagenic activation

of the compound (FEDRIP 1999). Dr. Harold F. Hemond at the Massachusetts Institute of Technology is

investigating the quantification of episodic and chronic releases of chromium from sediments into the

Aberjona watershed (FEDRIP 1999). Dr. Andrew J. Friedland at Dartmouth College is attempting to

establish the mobility rates of metals such as chromium in terrestrial ecosystems (FEDRIP 1999).

Dr. Friedland is using over 100 years of atmospheric deposition records and is evaluating potential release

rates of metals into surface water, aquatic ecosystems, and groundwater.

Remedial investigations and feasibility studies conducted at the NPL sites contaminated with chromium

will add to the available database on exposure levels in environmental media, exposure levels in humans,

and exposure registries, and will increase current knowledge regarding the transport and transformation of

chromium in the environment. No other long-term research studies regarding the environmental fate and

transport of chromium, or occupational or general population exposures to chromium were identified.

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